Technical Reports SeriEs No.
42I
Management of Waste
Containing Tritium
and Carbon-14
MANAGEMENT OF
WASTE CONTAINING
TRITIUM AND CARBON-14
The following States are Members of the International Atomic Energy Agency:
AFGHANISTAN
ALBANIA
ALGERIA
ANGOLA
ARGENTINA
ARMENIA
AUSTRALIA
AUSTRIA
AZERBAIJAN
BANGLADESH
BELARUS
BELGIUM
BENIN
BOLIVIA
BOSNIA AND HERZEGOVINA
BOTSWANA
BRAZIL
BULGARIA
BURKINA FASO
CAMEROON
CANADA
CENTRAL AFRICAN
REPUBLIC
CHILE
CHINA
COLOMBIA
COSTA RICA
CÔTE D’IVOIRE
CROATIA
CUBA
CYPRUS
CZECH REPUBLIC
DEMOCRATIC REPUBLIC
OF THE CONGO
DENMARK
DOMINICAN REPUBLIC
ECUADOR
EGYPT
EL SALVADOR
ERITREA
ESTONIA
ETHIOPIA
FINLAND
FRANCE
GABON
GEORGIA
GERMANY
GHANA
GREECE
GUATEMALA
HAITI
HOLY SEE
HONDURAS
HUNGARY
ICELAND
INDIA
INDONESIA
IRAN, ISLAMIC REPUBLIC OF
IRAQ
IRELAND
ISRAEL
ITALY
JAMAICA
JAPAN
JORDAN
KAZAKHSTAN
KENYA
KOREA, REPUBLIC OF
KUWAIT
KYRGYZSTAN
LATVIA
LEBANON
LIBERIA
LIBYAN ARAB JAMAHIRIYA
LIECHTENSTEIN
LITHUANIA
LUXEMBOURG
MADAGASCAR
MALAYSIA
MALI
MALTA
MARSHALL ISLANDS
MAURITIUS
MEXICO
MONACO
MONGOLIA
MOROCCO
MYANMAR
NAMIBIA
NETHERLANDS
NEW ZEALAND
NICARAGUA
NIGER
NIGERIA
NORWAY
PAKISTAN
PANAMA
PARAGUAY
PERU
PHILIPPINES
POLAND
PORTUGAL
QATAR
REPUBLIC OF MOLDOVA
ROMANIA
RUSSIAN FEDERATION
SAUDI ARABIA
SENEGAL
SERBIA AND MONTENEGRO
SEYCHELLES
SIERRA LEONE
SINGAPORE
SLOVAKIA
SLOVENIA
SOUTH AFRICA
SPAIN
SRI LANKA
SUDAN
SWEDEN
SWITZERLAND
SYRIAN ARAB REPUBLIC
TAJIKISTAN
THAILAND
THE FORMER YUGOSLAV
REPUBLIC OF MACEDONIA
TUNISIA
TURKEY
UGANDA
UKRAINE
UNITED ARAB EMIRATES
UNITED KINGDOM OF
GREAT BRITAIN AND
NORTHERN IRELAND
UNITED REPUBLIC
OF TANZANIA
UNITED STATES OF AMERICA
URUGUAY
UZBEKISTAN
VENEZUELA
VIETNAM
YEMEN
ZAMBIA
ZIMBABWE
The Agency’s Statute was approved on 23 October 1956 by the Conference on the Statute of
the IAEA held at United Nations Headquarters, New York; it entered into force on 29 July 1957.
The Headquarters of the Agency are situated in Vienna. Its principal objective is “to accelerate and
enlarge the contribution of atomic energy to peace, health and prosperity throughout the world’’.
© IAEA, 2004
Permission to reproduce or translate the information contained in this publication may be
obtained by writing to the International Atomic Energy Agency, Wagramer Strasse 5, P.O. Box 100,
A-1400 Vienna, Austria.
Printed by the IAEA in Austria
July 2004
STI/DOC/010/421
TECHNICAL REPORTS SERIES No. 421
MANAGEMENT OF
WASTE CONTAINING
TRITIUM AND CARBON-14
INTERNATIONAL ATOMIC ENERGY AGENCY
VIENNA, 2004
IAEA Library Cataloguing in Publication Data
Management of waste containing tritium and carbon-14.
— Vienna : International Atomic Energy Agency, 2004.
p. ; 24 cm. — (Technical reports series, ISSN 0074–1914 ; no. 421)
STI/DOC/010/421
ISBN 92–0–114303–6
Includes bibliographical references.
1. Radioactive waste disposal. 2. Tritium. 3. Carbon. 4. Hazardous
wastes — Management. 5. Environmental monitoring. I. International
Atomic Energy Agency. II. Technical reports series (International
Atomic Energy Agency) ; 421.
IAEAL
04–00360
FOREWORD
Carbon-14 and tritium are radioisotopes produced as a by-product or
special product in various nuclear reactor systems and globally in the
atmosphere by cosmic ray interaction with nitrogen and hydrogen, respectively.
Owing to their relatively long half-lives, high residence time in the
environment, high isotopic exchange rate and ease of assimilation into living
matter, it is necessary to control their production at nuclear facilities. There is
also a requirement for the proper management of related waste and material,
because of the potential impact on human health. The purpose of this report is
to review and analyse experience in the application of different organizational
and technological approaches to the management of waste containing 14C and
tritium. This report also reviews different sources of waste containing 14C and
tritium and their characteristics important in the selection of appropriate
methods for the processing, storage, disposal and release of this type of waste.
It is also intended by the publication of this report to update the information on
the management of tritium contaminated waste published by the IAEA in 1981
in Technical Reports Series No. 203, Handling of Tritium-Bearing Wastes, and,
in 1991, in Technical Reports Series No. 324, Safe Handling of Tritium: Review
of Data and Experience.
This report was prepared by experts from five countries through a series
of Consultants Meetings. The IAEA officer responsible for the preparation of
this report was V. Efremenkov of the Division of Nuclear Fuel Cycle and
Waste Technology. The IAEA is grateful to all experts who contributed to the
preparation of this report.
EDITORIAL NOTE
The use of particular designations of countries or territories does not imply any
judgement by the publisher, the IAEA, as to the legal status of such countries or territories,
of their authorities and institutions or of the delimitation of their boundaries.
The mention of names of specific companies or products (whether or not indicated
as registered) does not imply any intention to infringe proprietary rights, nor should it be
construed as an endorsement or recommendation on the part of the IAEA.
CONTENTS
1.
2.
3.
INTRODUCTION . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
1
1.1. Background . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
1.2. Objectives and scope . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
1
2
ENVIRONMENTAL AND REGULATORY ISSUES . . . . . . . . . .
3
2.1. Problems associated with 14C and tritium in waste . . . . . . . . . .
2.2. Regulatory issues . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
2.2.1. Argentina . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
2.2.2. Canada . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
2.2.3. France . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
2.2.4. Russian Federation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
2.2.5. United Kingdom . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
2.2.6. United States of America . . . . . . . . . . . . . . . . . . . . . . . . .
3
4
6
6
7
7
8
8
PRODUCTION AND EMISSION PATHWAYS . . . . . . . . . . . . . . .
11
3.1. Carbon-14 production and release . . . . . . . . . . . . . . . . . . . . . . .
3.1.1. Natural production in the atmosphere . . . . . . . . . . . . . .
3.1.2. Production in nuclear explosions . . . . . . . . . . . . . . . . . .
3.1.3. Production in and release from nuclear
power reactors . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
3.1.3.1. Light water reactors . . . . . . . . . . . . . . . . . . . . .
3.1.3.2. Heavy water reactors . . . . . . . . . . . . . . . . . . . .
3.1.3.3. Magnox reactors and advanced gas
cooled reactors . . . . . . . . . . . . . . . . . . . . . . . . . .
3.1.3.4. High temperature gas cooled reactors . . . . . .
3.1.3.5. Fast breeder reactors . . . . . . . . . . . . . . . . . . . .
3.1.3.6. Summary of waste containing 14C
produced by reactor operation . . . . . . . . . . . .
3.1.4. Release during spent fuel reprocessing . . . . . . . . . . . . .
3.2. Tritium production and release . . . . . . . . . . . . . . . . . . . . . . . . . .
3.2.1. Production and release in nuclear power reactors . . . .
3.2.1.1. Light water reactors . . . . . . . . . . . . . . . . . . . . .
3.2.1.2. Heavy water reactors . . . . . . . . . . . . . . . . . . . .
3.2.1.3. Other types of reactor . . . . . . . . . . . . . . . . . . . .
3.2.2. Release during spent fuel reprocessing . . . . . . . . . . . . .
11
11
11
11
14
18
21
23
24
24
25
28
29
31
31
32
33
4.
5.
3.3. Categories of waste containing
14
C and tritium . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
3.3.1. Reactor operation and fuel cycle waste . . . . . . . . . . . . .
3.3.2. Other waste . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
36
36
37
REDUCTION OF PRODUCTION AND RELEASE . . . . . . . . . .
37
4.1. Reduction of 14C production and release . . . . . . . . . . . . . . . . . .
4.2. Reduction of tritium production and release . . . . . . . . . . . . . . .
37
39
TECHNOLOGIES TO CAPTURE 14C AND TRITIUM FROM
EMISSION STREAMS . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
41
5.1. Removal of 14C from gas streams . . . . . . . . . . . . . . . . . . . . . . . .
5.1.1. Single-step chemical reaction involving absorption
in an alkaline earth hydroxide slurry or solid . . . . . . . .
5.1.1.1. Alkaline slurry scrubber . . . . . . . . . . . . . . . . . .
5.1.1.2. Alkaline packed bed column . . . . . . . . . . . . . .
5.1.2. Two-step chemical reaction involving sodium
hydroxide and lime slurry . . . . . . . . . . . . . . . . . . . . . . . .
5.1.2.1. Double alkali process . . . . . . . . . . . . . . . . . . . .
5.1.3. Physical absorption . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
5.1.3.1. Gas absorption by wet scrubbing . . . . . . . . . .
5.1.3.2. Ethanolamine scrubbing . . . . . . . . . . . . . . . . . .
5.1.3.3. Absorption in a fluorocarbon solvent . . . . . . .
5.1.4. Physical adsorption on an active surface . . . . . . . . . . . .
5.1.5. Other methods . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
5.2. Removal of 14C from liquid waste . . . . . . . . . . . . . . . . . . . . . . . .
5.3. Separation of tritium from spent fuel . . . . . . . . . . . . . . . . . . . . .
5.3.1. Voloxidation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
5.3.2. Pyrochemical processing . . . . . . . . . . . . . . . . . . . . . . . . .
5.4. Removal of tritium from gas streams . . . . . . . . . . . . . . . . . . . . .
5.4.1. Removal of HT . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
5.4.2. Removal of HTO . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
5.4.2.1. Molecular sieves . . . . . . . . . . . . . . . . . . . . . . . .
5.4.2.2. Dehumidification . . . . . . . . . . . . . . . . . . . . . . .
5.5. Removal of tritium from liquid waste . . . . . . . . . . . . . . . . . . . . .
5.5.1. Tritium enrichment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
5.5.2. Other tritium removal technologies . . . . . . . . . . . . . . . .
41
42
42
43
44
45
45
45
46
46
47
47
49
51
51
51
52
52
52
53
53
53
54
58
6.
7.
ANALYTICAL AND MONITORING METHODS . . . . . . . . . . . .
59
6.1. Carbon-14 monitoring systems . . . . . . . . . . . . . . . . . . . . . . . . . .
6.1.1. Carbon-14 sample collection . . . . . . . . . . . . . . . . . . . . . .
6.1.1.1. Air samples . . . . . . . . . . . . . . . . . . . . . . . . . . . .
6.1.1.2. Water samples . . . . . . . . . . . . . . . . . . . . . . . . . .
6.1.1.3. Vegetation and soils . . . . . . . . . . . . . . . . . . . . .
6.1.2. Carbon-14 sample preparation . . . . . . . . . . . . . . . . . . . .
6.1.3. Analytical methods for 14C . . . . . . . . . . . . . . . . . . . . . . .
6.1.3.1. Gas counting . . . . . . . . . . . . . . . . . . . . . . . . . . .
6.1.3.2. Liquid scintillation counting . . . . . . . . . . . . . .
6.1.3.3. Accelerator mass spectrometry . . . . . . . . . . . .
6.1.4. Carbon-14 monitoring methods . . . . . . . . . . . . . . . . . . .
6.2. Tritium monitoring systems . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
6.2.1. Air samples . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
6.2.2. Samples of liquids . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
6.2.3. Tritium analytical methods and monitoring systems . .
6.2.3.1. Ionization chambers . . . . . . . . . . . . . . . . . . . . .
6.2.3.2. Proportional counters . . . . . . . . . . . . . . . . . . . .
6.2.3.3. Scintillation crystal detectors . . . . . . . . . . . . . .
6.2.3.4. Mass spectrometers . . . . . . . . . . . . . . . . . . . . . .
6.2.3.5. Liquid scintillation counters . . . . . . . . . . . . . .
6.2.3.6. Portable room air monitors . . . . . . . . . . . . . . .
6.2.3.7. Fixed station room air monitors . . . . . . . . . . .
6.2.3.8. Glovebox atmosphere monitors . . . . . . . . . . .
6.2.3.9. Hood and exhaust duct air monitors . . . . . . . .
6.2.3.10. Exhaust stack air monitors . . . . . . . . . . . . . . . .
6.2.4. Specialized instrumentation . . . . . . . . . . . . . . . . . . . . . . .
6.2.4.1. Remote field tritium analysis system . . . . . . .
6.2.4.2. Surface activity monitor . . . . . . . . . . . . . . . . . .
6.2.4.3. Breathalyser . . . . . . . . . . . . . . . . . . . . . . . . . . . .
60
60
60
61
61
61
63
63
63
63
63
64
64
66
67
67
67
68
68
68
69
69
69
69
70
70
70
70
71
IMMOBILIZATION AND WASTE FORM EVALUATION . . . .
71
7.1. Immobilization technologies for waste containing 14C . . . . . . .
7.1.1. Cementation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
7.1.2. Other methods . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
7.2. Immobilization technologies for waste containing tritium . . . .
7.2.1. Drying agents . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
7.2.2. Hydraulic cements . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
7.2.3. Organic agents . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
72
73
73
74
75
75
76
8.
9.
7.2.4. Hydrides . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
7.3. Assessment of immobilized waste form
and quality control . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
7.3.1. Waste form characterization . . . . . . . . . . . . . . . . . . . . . .
7.3.2. Waste form testing . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
76
STORAGE AND DISPOSAL . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
81
8.1. Waste acceptance requirements . . . . . . . . . . . . . . . . . . . . . . . . . .
8.1.1. France . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
8.1.2. Japan . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
8.1.3. Spain . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
8.1.4. United Kingdom . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
8.1.5. United States of America . . . . . . . . . . . . . . . . . . . . . . . . .
8.2. Storage and disposal options for waste
containing 14C and tritium . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
8.2.1. Storage options . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
8.2.1.1. Storage of HTO waste . . . . . . . . . . . . . . . . . . .
8.2.1.2. Storage of tritium gas . . . . . . . . . . . . . . . . . . . .
8.2.1.3. Storage in engineered structures . . . . . . . . . . .
8.2.1.4. Storage of irradiated graphite waste . . . . . . . .
8.2.2. Disposal options . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
8.2.2.1. Near surface disposal . . . . . . . . . . . . . . . . . . . .
8.2.2.2. Liquid injection into geological formations . .
8.2.2.3. Disposal in geological formations . . . . . . . . . .
8.2.3. Other options . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
8.3. Strategy for the management of waste containing
14
C and tritium . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
81
83
84
85
86
86
CONCLUSIONS . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
77
78
79
87
87
88
89
89
90
91
91
92
93
93
94
94
REFERENCES . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 97
CONTRIBUTORS TO DRAFTING AND REVIEW . . . . . . . . . . . . . . . 109
1. INTRODUCTION
1.1. BACKGROUND
Carbon-14 and tritium are naturally occurring radioisotopes produced
continuously in the atmosphere by cosmic ray neutron interaction with
nitrogen and hydrogen, respectively, and are also produced as a by-product or
special product in nuclear reactor systems. Carbon-14 and tritium have rather
long half-lives (tritium = 12.3 a and 14C = 5730 a) and residence times in the
atmosphere and hydrosphere sufficient for transport processes to distribute
them worldwide. Their radiological impact is therefore not limited to the region
of release, and they may be distributed globally in a nearly uniform manner.
Carbon-14 and tritium are weak beta emitters and do not present an external
radiation hazard. Owing to their long half-lives, high isotopic exchange rates
and ease of incorporation into living organisms, however, it is important to
control their release from nuclear facilities and waste management sites to the
environment by the use of appropriate waste management strategies and
practices.
Carbon-14 in the nuclear fuel cycle is produced by neutron interaction
with 13C, 14N, 15N, 16O and 17O, which may be present in the nuclear fuels and
the moderator and primary coolant systems of nuclear reactors. Tritium is
produced in nuclear power reactors during the fission of heavy nuclei and by
neutron interaction with coolants, moderators and some light elements, such as
lithium, beryllium and boron. The rates of production of 14C and tritium are
dependent on the type of the reactor and its capacity.
The amounts of 14C and tritium in waste generated by the nuclear
industry vary as a function of, for example, reactor type, plant operation, the
method of fuel reprocessing and the radioisotope production process. One
common property of both is that they are difficult to assay, since they are weak
beta emitters and do not emit gamma radiation. These factors and some other
properties mean that, in the context of the management of radioactive waste,
14
C and tritium are considered to be problematic and that their properties need
to be specifically considered when selecting methods for the treatment, conditioning and packaging of waste containing 14C and tritium and when establishing release limits and waste acceptance requirements (WARs) for these
particular radioisotopes.
1
1.2. OBJECTIVES AND SCOPE
The primary objective of this report is to provide Member States and
responsible organizations with information on the organizational principles
and technical options for the management of radioactive waste and effluents
containing 14C and tritium, including waste collection, separation, treatment,
conditioning, and storage and/or disposal. A subsidiary objective of this report
is to identify areas in which a lack of knowledge or proven technology limits
achievement of the primary objective. These objectives are achieved by
reviewing the different sources and characteristics of waste streams containing
14
C and tritium and by analysing methods for the processing, storage and/or
disposal of these types of waste, both well proven methods and those at an
advanced stage of development.
This report should aid the reader in the selection of an appropriate
management strategy for waste and effluents containing 14C and tritium, taking
into account the special requirements and limitations associated with the
disposal of this type of waste and the release of these radioisotopes. To address
the identified objectives, this report:
(a)
(b)
(c)
(d)
(e)
(f)
2
Overviews some existing regulations for emissions of 14C and tritium and
the storage and/or disposal of waste containing these radioisotopes.
Identifies the nature and magnitude of the problems pertaining to waste
containing 14C and tritium by analysing sources, production rates and
potential emission pathways to the environment of these radioisotopes in
various nuclear operations, such as reactor operations and spent nuclear
fuel reprocessing.
Identifies and describes the characteristics, physical and chemical forms,
and properties of waste containing 14C and tritium produced by nuclear
facilities.
Analyses approaches applied for the reduction and/or minimization of
14
C and tritium production in nuclear operations; this should be achieved
by analysing the different factors influencing the generation of these
radionuclides and by taking into account the process chemistry and
mechanism of their release.
Reviews technologies for the removal of 14C and tritium from liquid and
gaseous waste streams and for concentrating the levels of 14C and tritium
in such streams, and discusses their merits and limitations.
Reviews sampling methods, procedures, analytical techniques and
equipment for analysing and monitoring these radioisotopes in waste
streams and effluents.
(g)
(h)
(i)
Reviews options and specific technologies for the immobilization and/or
stabilization of waste containing 14C and tritium, and discusses the aspects
of waste form evaluation in respect of the safety requirements, the
characteristics of the waste forms and regulatory, administrative and
technological issues.
Discusses specific requirements and practices for the storage and/or
disposal of waste containing 14C and tritium based on existing national
and international regulations.
Analyses trends in the development and application of novel technical
approaches in the management of waste containing 14C and tritium.
2. ENVIRONMENTAL AND REGULATORY ISSUES
2.1. PROBLEMS ASSOCIATED WITH 14C AND TRITIUM IN WASTE
Carbon-14 is a radioactive isotope of carbon and is a pure beta emitter
with a half-life of 5730 years; it decays to 14N by emitting low energy beta
radiation with an average energy of 49.5 keV and a maximum energy of
156 keV. Carbon-14 is easily transferred during biological processes and soil–
plant interactions involving carbon compounds. The metabolism and kinetics of
14
C in the human body follow those of ordinary carbon. Inhaled 14CO2 rapidly
equilibrates with the air in the lungs and enters many components of body
tissue. The biological half-life of 14C is approximately 40 days. It has been found
that accumulation of 14C in the human body via respiration is insignificant
compared with that from ingestion of contaminated food. In addition, 14C can
be easily concentrated in the food chain. Studies have shown concentration
factors of 5000 for fish and molluscs and 2000 for soil sediments.
Tritium is a radioactive isotope of hydrogen and has a half-life of 12.3
years; it decays to 3He by emitting low energy beta radiation with an average
energy of 5.7 keV and a maximum energy of 18.6 keV. Transfer of environmental tritium oxide (or tritiated water) (HTO) to humans takes place via
inhalation, diffusion through the skin and ingestion. Inhalation is the only
meaningful pathway of tritium gas (containing both hydrogen and tritium HT)
to humans.
Radioactive waste containing 14C and tritium is continuously generated
by the nuclear industry in, for example, nuclear reactor operations, spent fuel
reprocessing and radioisotope production, and in medical research. Both 14C
3
and tritium are considered to be mobile in the environment, and specific
considerations are therefore required for the management of waste containing
these radioisotopes; for example, the disposal of waste containing significant
amounts of 14C will be more difficult than that of waste containing only short
lived nuclides, owing to the stringent WARs for 14C. The minimization and
segregation of waste containing long lived nuclides (e.g. 14C) are therefore
important factors for an effective waste management approach.
2.2. REGULATORY ISSUES
The objective of the management of waste containing 14C and tritium is to
protect human health and the environment from the harm associated with
these contaminants due to direct exposure and bioaccumulation in the food
chain. Environmental protection includes the protection of living organisms
other than humans and the protection of natural resources, including land,
forests, water and raw materials, together with a consideration of nonradiological environmental impacts.
In assessing the impact of 14C and tritium releases, the possible exposure
of an individual in the immediate vicinity outside an exclusion boundary area,
and of the public at large, should be considered. Appropriate programmes on
effluent and environmental monitoring are therefore required for each nuclear
facility to ensure protection of the public from radioactive discharges.
In general, the owners or operators of nuclear facilities are responsible
for setting up and implementing the technical and organizational measures
necessary for ensuring protection of the public from radioactive discharges. In
particular, they are responsible for implementing any conditions or limitations
specified by the regulatory body in an authorization [1]. This authorization can
be in the form of a registration (for practices with low to moderate associated
risks to the public), a licence (for practices with high potential risks to the
public) or a similar document.
The application for an authorization such as a licence should contain the
following information:
(a)
(b)
An appropriate safety assessment, including an explanation of how radiological protection has been optimized;
An assessment of the nature, magnitude and likelihood of exposures
attributed to the discharges.
In addition, the application should also address the issues of waste
generation and management interdependences. In general, discharge limits for
4
the radionuclides concerned would be included in, or would accompany, a
licence issued by the regulatory body that allows the operation of the facility.
Authorized discharge limits are set by the regulatory body. The discharge
limits discussed in this report are restricted to discharges to the environment of
radioactive substances (i.e. 14C and tritium) in the form of airborne (gases and
aerosols) or liquid (to surface water bodies only) effluents from the normal
operation of practices and sources within practices. The sources considered
range from radionuclides used for medical and research purposes to nuclear
reactors and spent fuel reprocessing facilities.
General principles for setting limits for the release of radionuclides from
nuclear facilities have been provided by the IAEA in several publications
[1–3]. The discharge limits should satisfy the requirements for the optimization
of protection and the condition that doses to the critical group should not
exceed the appropriate dose constraints. They should also reflect the requirements of a well designed and well managed practice and should provide a
margin for operational flexibility and variability [1].
Discharge limits are usually attached to or incorporated into the facility
licence and become the legal limits with which the operator or licensee should
comply. Discharge authorizations are normally set in terms of annual limits.
While these are the primary limits, shorter term limits can be set in order to
trigger investigations and ensure that the procedure used and the associated
conditions and assumptions used to estimate dose limits remain valid (e.g. to
prevent significantly higher doses being received due to higher than normal
discharges in conditions of poor dispersal in the environment) [1].
In order to evaluate releases of 14C from nuclear facilities, a background
activity of 250 Bq of 14C per kilogram of stable carbon has been used by
regulatory bodies such as the Institut de protection et de sûreté nucléaire in
France, the National Radiological Protection Board in the United Kingdom
and safety authorities in Switzerland. Any 14C level above this background
level, other than from the normal production of 14C by cosmic ray neutron
interaction with nitrogen in the atmosphere, will be considered to be pollution.
Discharge limits for radioisotopes are established in most countries in
accordance with the recommendations of the International Commission on
Radiological Protection (ICRP) [4]. These limits differ from one site to another
depending on assumptions made on the nature of the effluent and on the
environment into which the discharges are made. Some examples of the
approaches used in various countries to determine the discharge limits for 14C
and tritium are summarized in Sections 2.2.1–2.2.6.
5
2.2.1.
Argentina
The regulatory authorities in Argentina have specified two dose
constraints on releases of radioactive material from nuclear power plants to the
environment [5]. These are that the:
(a)
(b)
Individual annual dose in the critical group resulting from the release
shall not exceed 0.3 mSv/a;
Collective dose commitment per unit of practice shall not exceed
1.5 × 10–2 Sv·MW(e)–1·a–1 of electric energy generated by the plant.
For example, the derived release limits (DRLs) for airborne and
waterborne tritium from the Embalse nuclear power plant (a CANDU heavy
water reactor (HWR)) are 3.7 × 104 TBq/a and 3.7 × 103 TBq/a, respectively.
These DRLs result in a 0.3 mSv/a dose to a member of the public critical group.
2.2.2.
Canada
DRLs for the release of radionuclides from nuclear facilities in Canada
are set to ensure that the committed dose to a critical group from one year’s
release does not exceed the limit of the annual dose to a member of the public
set by the Canadian Nuclear Safety Commission (formerly the Atomic Energy
Control Board) [6]. The DRL is defined as the radioactive release over a year
that would expose members of the critical group to the regulatory dose limit. In
general, nuclear power plants set their operating targets for releases of each
radionuclide below 1% of the DRL. The annual and weekly DRLs for airborne
14
C and tritium releases from two CANDU 600 MW(e) nuclear power plants in
Canada are shown in Table 1 [6].
The large variations in the calculated DRLs based on a public dose limit
of 5 mSv/a for various sites arise primarily owing to the location of the nuclear
reactor site and the proximity of a local population outside the exclusion
boundary area. The Gentilly 2 site is located near to a farm and it is considered
that a member of the critical group consumes water from the St. Lawrence
River 3 km downstream of the station. In contrast, the Point Lepreau nuclear
power plant is located in a relatively remote area. As a result, the DRLs for
airborne releases from the Point Lepreau nuclear power plant tend to be higher
than those from Gentilly 2.
6
TABLE 1. DRLs FOR AIRBORNE TRITIUM AND 14C DISCHARGES
FROM CANDU 600 MW(e) NUCLEAR POWER PLANTS IN CANADA
DRLs for airborne
tritium discharges
DRLs for airborne
14
C discharges
TBq/week
TBq/a
TBq/week
TBq/a
Point Lepreau
3.00 × 104
1.56 × 106
3.00 × 102
1.56 × 104
Gentilly 2
8.50 × 103
4.42 × 105
1.70 × 101
8.84 × 102
2.2.3.
France
In France the limits to discharges of gaseous and liquid effluents from
nuclear power plants (mainly pressurized water reactor (PWR) plants) and fuel
reprocessing plants imposed by the regulatory authority are relevant for
emissions of aerosols, halogens and gross beta–gamma activities; for example,
the discharge limits of tritium in gaseous and liquid effluents are 5.6 TBq/a
(150 Ci/a) and 56 TBq/a (1500 Ci/a), respectively, from two nuclear power units
with a 900 MW(e) capacity each. The discharge limits of tritium from the La
Hague reprocessing plant, which consists of two units (UP2 and UP3) with a
total processing capacity of 1700 t of heavy metal per year, are 2200 TBq/a
(59 460 Ci/a) and 37 000 TBq/a (1 000 000 Ci/a) for gaseous and liquid effluents,
respectively.
In general, the average annual emissions of tritium from nuclear power
plants are approximately 40% of the airborne discharge limit and 35% of the
waterborne discharge limit set by the regulatory authority. For the La Hague
reprocessing plant, about 4% and 37% of the discharge limits for gaseous and
liquid effluents, respectively, are released per year.
At present in France there are no specific limits set for the release of 14C
from nuclear facilities, except the discharge limits for gross beta–gamma
emissions, although such discharge limits are likely to be imposed in the near
future.
2.2.4.
Russian Federation
Standards on radiological protection in the Russian Federation are
essentially based on the International Basic Safety Standards for Protection
against Ionizing Radiation and for the Safety of Radiation Sources [3] and on
the recommendations of the ICRP [4].
According to the Russian radiation safety standards [7], the committed
effective dose incurred by any member of the public should not exceed
7
1 mSv/a, and that for a worker during one year’s performance of a practice (e.g.
reactor operation) is 20 mSv. The permissible intake values and corresponding
concentrations of 14C and tritium to ensure adequate human radiological
protection, based on the assumptions that a man (operator) breathes on
average 1.2 m3 of air in an hour and a member of the public consumes 2 kg of
water per day, are presented in Tables 2 and 3.
The discharge limits for 14C and tritium are established taking into
account the combined effect of all radionuclides present in released gas and/or
liquid effluents. A safety factor as high as five is ordinarily used in setting
discharge limits for operating existing nuclear power plants, while for newly
designed nuclear power plants a safety factor of 25 is usually employed.
2.2.5.
United Kingdom
The Environment Agency in the UK has revised the discharge limits for
gaseous emissions from Magnox nuclear power plants operated by British
Nuclear Fuels (BNFL) in accordance with the present status of the plant (i.e. in
operation or in a decommissioning phase). Table 4 shows the existing and
drafted revised emission limits for 14C and tritium at various Magnox nuclear
power plants in the UK [8].
The revised limits are determined based on the dose uptake by the critical
group. All dose limits are set much lower than the recommended maximal
limits defined in ICRP 60 [4] for the public of 1 mSv/a. It should be noted that
the dose limit set for the Trawsfynnyd plant is relatively low because the plant
is in the decommissioning phase.
2.2.6.
United States of America
The Nuclear Regulatory Commission (NRC) and the United States
Department of Energy (USDOE) have the primary responsibility in the USA
for regulating nuclear facilities. The NRC regulates and licenses nuclear power
reactors, non-power research reactors and nuclear material (including related
waste). There are 32 states that have formal agreements with the NRC by
which those states have assumed regulatory responsibility over certain byproduct, source and small quantities of special nuclear material. The USDOE
oversees and monitors contractors that operate federally owned facilities that
produce or process nuclear material and nuclear waste. Both NRC and
USDOE regulated facilities are typically subject to all federal regulations
governing nuclear facilities.
The Environmental Protection Agency (EPA) defines national policies,
goals and regulations concerning the environment. The National Emission
8
TABLE 2. LIMITATIONS ON TRITIUM AND 14C INTAKE FOR
NUCLEAR OPERATORS (RUSSIAN FEDERATION)
Type of inhaled
nuclide (chemical
form)
Dose coefficient
(Sv/Bq)
Upper annual
intake limit
(Bq/a)
Permissible
specific activity
(Bq/m3)
Tritium
HTO vapour
1.8 × 10–11
1.1 × 1090
4.4 × 105
Tritium
Gaseous tritium
1.8 × 10–15
1.1 × 1013
4.4 × 109
–13
11
4.4 × 107
Tritium
Tritiated methane
1.8 × 10
1.1 × 10
14
Elementary carbon
5.8 × 10–10
3.4 × 1070
1.4 × 104
14
CO2
6.2 × 10–12
3.2 × 1090
6.2 × 106
14
CO
8.0 × 10–13
2.5 × 1010
1.0 × 107
C
C
C
TABLE 3. LIMITATIONS ON TRITIUM AND 14C INTAKE FOR THE
GENERAL PUBLIC (RUSSIAN FEDERATION)
Type of inhaled
nuclide (chemical
form)
Dose coefficient
(Sv/Bq)
Upper annual
intake limit
(Bq/a)
Permissible
specific activity
(Bq/m3)
3.7 × 106
1.9 × 103
Intake with air
Tritium
All types
2.7 × 10–10
Intake with water and food
Tritium
Inorganic
compounds
4.8 × 10–11
2.1 × 107
7.7 × 103
Tritium
Organic
compounds
1.2 × 10–10
8.3 × 106
3.3 × 103
14
All types
2.5 × 10–90
4.0 × 105
5.5 × 101
Intake with air
C
Intake with water and food
14
C
All types
1.6 × 10–90
6.3 × 105
2.4 × 102
Standards for Hazardous Air Pollutants (NESHAPS) require that the off-site
dose from airborne effluents be less than 100 mSv/a (10 mrem/a) for all radionuclides combined.
9
TABLE 4. AIRBORNE TRITIUM AND 14C DISCHARGE LIMITS AT
VARIOUS MAGNOX NUCLEAR POWER PLANTS IN THE UK
Bradwell
Dungeness Hinkley
A
Point A
Oldbury
Wylfa
Sizewell
A
Trawsfynnyd
Tritium
Existing limit
(TBq/a)
3
25
5
7
73.5
10
20
Limit in the draft
authorization
(TBq/a)
1.5
2.6
5
9
18
3.5
0.75
0.26
0.14
0.48
1.2
1
0.51
0.066
Critical group
dose from the
draft
authorization
limit (µSv/a)
14
C
Existing limit
(TBq/a)
5
4
6
1.5
24.7
5
2.4
Limit in the draft
authorization
(TBq/a)
0.6
5
4
5
2.3
2
0.01
14
12
50
86
17
38
0.15
80
165
116
191
64
174
0.22
Critical group
dose from the
draft
authorization
limit (µSv/a)
Total dose to the
critical group
(µSv/a)
For beta–gamma emitters in water, the standard is derived from average
annual concentrations calculated to produce a total body or organ dose of
40 mSv/a (4 mrem/a). The EPA safe drinking water standard for 14C is 74 Bq/L
(2000 pCi/L) and that for tritium is 740 Bq/L (20 000 pCi/L). Note that these
are standards for drinking water and not discharge limits, but they may be used
by the EPA, other federal regulators and state regulators in setting specific
discharge limits [9].
10
3. PRODUCTION AND EMISSION PATHWAYS
3.1. CARBON-14 PRODUCTION AND RELEASE
3.1.1.
Natural production in the atmosphere
Natural 14C is produced in the upper atmosphere by the 14N(n,p)14C
reaction induced by cosmic ray neutrons. The annual production rate by this
mechanism is estimated to be 1.4 × 106 GBq (3.8 × 104 Ci), with a total
inventory of 1.4 × 108 GBq (3.8 × 106 Ci) in the atmosphere. A much larger
quantity of 14C (approximately 1.0 × 1010 GBq) is located in the deep oceans
and exchanges with atmospheric carbon [6].
3.1.2.
Production in nuclear explosions
Carbon-14 is formed in nuclear explosions as a result of neutron capture
on nitrogen and resides in the atmosphere as 14CO2 [6]. The amount of
14
C added to the atmosphere and labile biosphere by atmospheric nuclear
weapon testing in the 1950s and 1960s has been estimated to be 2.2 × 108 GBq
(6.0 × 106 Ci) [10]. Figure 1 [11] shows 14C concentrations in the atmosphere
between 1955 and 1994. The high concentrations of 14C during the 1960s were
the result of atmospheric nuclear weapon testing.
3.1.3.
Production in and release from nuclear power reactors
The normal operation of nuclear reactors for the generation of electric
power produces various radioisotopes by fission within the fuel or by neutron
activation in the structural materials and component systems of the reactor.
The escape of these radioisotopes from the reactor and its auxiliary process
systems generates a variety of solid, liquid and gaseous radioactive waste.
Although the design of the reactor ensures that releases of liquid and airborne
waste are minimized, small quantities of radionuclides escape the systems and
are continuously discharged in various effluents. Carbon-14 is one of these
radionuclides.
The major 14C producing neutron activation reactions in nuclear power
reactors are:
(a)
The 14N(n,p)14C reaction with a very high thermal neutron capture crosssection (1.82 barn (1 barn = 10–24 cm2));
11
(b)
(c)
(d)
(e)
The 17O(n,a)14C reaction with a high thermal neutron capture crosssection (0.24 barn);
The 13C(n,g)14C reaction with a low cross-section (0.9 × 10–3 barn);
The 15N(n,d)14C reaction with a very low cross-section (2.5 × 10–7 barn);
The 16O(n,3He)14C reaction with a very low cross-section (5.0 × 10–8
barn).
In general, 14C is produced in nuclear power reactors by 14N(n,p)14C
reactions with nitrogen in fuels, moderators and coolants as a primary impurity,
by 17O(n,a)14C reactions in oxide fuels, moderators and coolants, and by
13
C(n,g)14C reactions in graphite moderators. Reactions (a), (b) and (c) are the
most important contributors to 14C production. Reactions (d) and (e) are
unimportant in thermal reactors. Carbon-14 is also a ternary fission product,
but the amount produced in this way is negligible.
The substrate atoms for the activation reactions (i.e. nitrogen, oxygen and
carbon) occur widely in fuel, and in cladding, moderator, coolant or structural
material, either as major constituents or as impurities. In consequence, 14C
produced in a nuclear power reactor can be released directly to the
environment from the coolant and/or moderator in a gaseous form or in much
220
Atmospheric CO2: northern hemisphere
Carbon-14 concentration (pmc)
200
Atmospheric CO2: southern hemisphere
180
160
Biosphere (15 years)
140
120
100
Surface ocean
80
1955
1960
1965
1970
1975
1980
1985
1990
Year
FIG. 1. Carbon-14 concentration in the atmosphere: 1955–1994.
12
1995
smaller quantities as liquid effluents. Carbon-14 can remain in the reactor core
until a reactor is decommissioned (e.g. in the graphite moderator of advanced
gas cooled reactors (AGR) and Magnox reactors) or it can pass to a fuel
reprocessing plant in the spent fuel and other fuel element components. The
14
C from the fuel will be released at the reprocessing plant into the off-gases at
the dissolution stage, and the cladding will constitute a separate solid waste
arising. Both the amounts of 14C produced and the chemical forms of 14C
present depend on the details of the reactor system, as does the subsequent
behaviour of the 14C and the pathways by which it can be released to the
environment [12].
The amounts of 14C produced by the various types of reactor vary considerably, depending on the fuel enrichment, temperature and relative masses of
the fuel, moderator and coolant, and on the concentrations of nitrogen
impurities in these systems. The production rate of 14C in a reactor can be
calculated from [13]:
A=
(
)
fmL
dN
sf 1 − e− λti
= Nsf − Nsfe− λti =
dt
M
where
A
f
L
σ
φ
λ
ti
N
m
M
is the activity produced (disintegrations per second);
is the fractional isotopic abundance of the target element (i.e. the
substrate element);
is Avogadro’s number (6.025 × 1023 mol–1);
is the thermal neutron cross-section in barns (10–24 cm2);
is the neutron flux in n·cm–2·s–1;
is the decay constant of 14C per second (ln 2/half-life);
is the irradiation time in seconds;
is the number of target atoms;
is the mass of the target element in grams;
is the atomic weight of the target element in g/mol.
A summary of typical 14C production rates in various types of reactor is
given in Table 5 [10].
Limited data on 14C emissions from nuclear power reactors have been
published. The release levels of 14C from nuclear reactors depend mainly on the
reactor type, its design and the site specific effluent treatment programmes in
place for the plant; for example, in light water reactors (LWRs) the 14C
produced in the moderators and coolants can be assumed to be essentially
released from the reactor to the environment. However, in HWRs more than
13
TABLE 5. CALCULATED 14C PRODUCTION RATES (GBq·GW(e)–1·a–1)
FOR VARIOUS TYPES OF REACTOR
Fuel
Fuel cladding
Coolant and
moderator
Graphite
moderator
Total
LWR–PWR
480
740
260
—
1 480
LWR–BWR
470
630
190
—
1 290
HWR
1 465
1 260
7 400
—
10 125
GCR–MGR
4 835
1 300
310
10 730
17 175
GCR–AGR
620
1 180
300
3 480
5 580
GCR–HTGR
190
—
1
3 180
3 371
FBR
200
300
—
500
—
Note: LWR: light water reactor; PWR: pressurized water reactor; BWR: boiling water
reactor; HWR: heavy water reactor; GCR: gas cooled reactor; MGR: Magnox reactor;
AGR: advanced gas cooled reactor; HTGR: high temperature gas cooled reactor; FBR:
fast breeder reactor.
half the 14C produced in the moderators and coolants is retained on system
purification ion exchange resins. Table 6 shows the estimated net release rates
of various types of reactor and reprocessing plants in 1998 [14]. The majority of
14
C released from reactors and fuel reprocessing plants is contained in airborne
effluents, and only a small amount is in the form of liquid effluents from
nuclear facilities.
For all types of reactor except PWRs, 14C is emitted mainly as CO2. For
boiling water reactors (BWRs) 80–95% of the released 14C appears to be CO2
and 5–20% hydrocarbons. For HWRs approximately 80% and 20% of the total
14
C released are CO2 and hydrocarbons, respectively. In comparison, the
airborne 14C released from PWRs is predominantly hydrocarbons (75–95%),
mainly methane, with only a small fraction in the form of CO2 (Table 7).
The mechanisms of 14C production and 14C release pathways in various
types of reactor are summarized in Sections 3.1.3.1–3.1.3.5.
3.1.3.1. Light water reactors
LWRs are widely used throughout the world and hence have a significant
effect on overall 14C emissions. Production of 14C depends on the enrichment of
the fuel, the relative mass of the fuel and moderator, the concentrations of
nitrogen impurities in the fuel and structural materials, and the temperatures of
14
TABLE 6. ESTIMATED 14C ANNUAL RELEASE RATES IN THE
ABSENCE OF CONTROL FROM REACTORS AND REPROCESSING
PLANTS
Number of units
Estimated 14C release
rate (GBq/a)
Total (GBq/a)
PWR
207
185
38 295
BWR
93
295
27 435
HWR (CANDU)
35
2 590
90 650
Graphite reactor
35
555
19 425
RBMK reactor
14
1 850
25 900
WWER reactor
47
1 850
86 950
Subtotal
431
Reprocessing plantsa
288 655
3
18 500
55 500
Total
344 155
Note: RBMK: high power channel type reactor; WWER: water cooled, water
moderated power reactor.
a
The reprocessing plants taken into account are La Hague (COGEMA, France), Sellafield (BNFL, UK) and Chelyabinsk (Russian Federation).
TABLE 7. COMPARISON OF THE CHEMICAL FORMS (RELATIVE
PERCENTAGE OF EACH SPECIES) OF 14C IN AIRBORNE RELEASES
FROM VARIOUS TYPES OF REACTOR
14
HWR (Bruce unit 7, Canada)a
CO2
14
CO
14
C hydrocarbons
65.5–72.8
0.2–3.7
26.7–34.4
HWR (Gentilly 2, Canada)
77.9–97.5
0.01–0.09
25.0–22.0
PWR (USA and Europe)
5–25.0
—
75–95 (CH4 and C2H6)
BWR (USA and Europe)
80–95.0
—
5–20
b
a
b
Data were measured by Atomic Energy of Canada Limited in November 1994.
Data were measured by Atomic Energy of Canada Limited in September and October
1995.
15
the fuel and moderator. Since there are considerable variations in design, very
precise calculations of 14C production in LWRs are probably not justified.
Calculations of 14C production rates in the fuel and coolant and/or
moderator of LWRs have been reported. The calculated values were highly
dependent on the assumptions made for the nitrogen impurities and 17O levels
in the fuel and the coolant and/or moderator. The ranges of calculated 14C
production rates in LWRs based on 25 ppm of nitrogen impurity in the fuel are
summarized in Table 8, and those in the cladding and structural materials are
given in Table 9 [12].
It can be assumed that most stainless steel structural materials remain
inside the reactor when fuel elements are removed and constitute a decommissioning waste. In an LWR the zircaloy cladding is the dominant source of 14C,
and contains approximately 50–60% of its total.
Carbon-14 is produced in the fuel and coolant and is distributed wherever
gas or fluid streams flow in the power plant. Leakage of plant systems allows
for eventual release to the environment, so the partitioning of original 14C in
various pathways is an important guide to the establishment of control
measures. A simplified diagram of flows of 14C in an LWR is shown in Fig. 2.
Carbon-14 is always carried by stable carbon compounds. In a BWR the
air entrained in the coolant is ejected from the main condenser. This off-gas is
fundamentally air, and therefore carbon, as CO2, exists in the similar ratio to
other constituents as it does in air. A very small amount of the stable carbon
remains in the coolant; this level is probably controlled by coolant chemistry. A
part of the 14C remains dissolved in the primary water purification and
treatment systems, causing smaller sources of release, for example in the
auxiliary building and finally in the active liquid waste processing system [15].
Water circulation
Fuel
LWR
Cladding
Reactor components
Reprocessing plant
Decommissioning
Absorption from off-gas
Coolant
Leakage
Immobilization
Discharge
Disposal
Environment
14
FIG. 2. Simplified diagram of C flows in an LWR reactor and at the fuel reprocessing
stage.
16
TABLE 8. RANGES OF CALCULATED 14C PRODUCTION RATES
(GBq·GW(e)–1·a–1) IN LWRs BASED ON 25 ppm OF NITROGEN
IMPURITY IN THE FUEL
Fuel
Coolant
Total
BWR
600–2340
190–420
790–2760
PWR
600–9000
110–230
710–1130
TABLE 9. CALCULATED 14C PRODUCTION RATES (GBq·GW(e)–1·a–1)
IN THE CLADDING AND STRUCTURAL MATERIALS OF LWRs
304 stainless
steel
302 stainless
steel
Zircaloy 2
Nicobraze 50
Total
BWR
1280–2040
—
630
—
1910–2670
PWR
0695–1110
177–125
350–740
4
1126–1979
The following systems have been considered to be release pathways for
gaseous 14C from a BWR [13]:
(a)
(b)
(c)
(d)
(e)
The condenser steam jet air ejector;
The turbine gland seal condenser exhaust;
The reactor building purge exhaust;
The turbine building ventilation system exhaust;
The radioactive waste building ventilation system exhaust.
The condenser steam jet air ejector is expected to be the most significant
release point (>99% of total 14C release).
In a PWR a portion of CO2 remains dissolved in the coolant, while most
leaks to the airspace in the reactor, where it is diluted in nitrogen. Dilution in
the off-gas streams is great enough to ensure an air-like composition with
respect to CO2. Owing to boric acid addition and buffering, a small amount of
CO2 is expected to remain in the primary coolant. A part of the 14C compounds
remains dissolved in the water and is released at different steps of the active
liquid waste treatment. Corrosion control for the secondary systems maintains
a low CO2 level, causing most to be released as gas in the air ejector. The
following systems in a PWR have been considered to be release pathways for
gaseous 14C [13]:
17
(a)
(b)
(c)
(d)
(e)
(f)
(g)
Primary off-gas treatment vents;
The steam generator blowdown tank vent exhaust;
The turbine gland seal condenser exhaust;
The fuel handling building ventilation exhaust;
The containment ventilation system exhaust;
The auxiliary building ventilation system exhaust;
The turbine building ventilation system exhaust.
The most significant element contributing to 14C release is the off-gas
stream from the primary off-gas treatment system, which accounts for approximately 70% of the total release.
The average normalized 14C releases from PWRs and BWRs are
summarized in Table 10. The typical average 14C release rate from an LWR is
approximately 185–370 GBq/a [11].
The total environmental release of 14C from the reactor, expressed as a
function of the production rate, is on average about 50% in BWRs and 30% in
PWRs [15].
3.1.3.2. Heavy water reactors
Carbon-14 in HWRs is produced mainly in four systems [15]:
(a)
(b)
(c)
(d)
The moderator system;
The heat transport (coolant) system;
Fuel elements;
The annulus gas system.
In general, 14C production in HWRs will be higher than in LWRs, owing
to the large quantity of heavy water (with 17O) in the moderator system.
Table 11 gives a breakdown of 14C production by a typical CANDU 600 MW(e)
nuclear reactor [6]. The majority of the 14C is produced in the heavy water
moderator.
The results of 14C distribution within a reactor system, modelled for the
CANDU system, are summarized in Table 12 [6]. The model predicts that 3.9%
of total 14C production is released into the atmosphere from the moderator
cover gas and the annulus gas systems. Approximately 3.4% of the 14C remains
in the fuel and 92.7% is removed by ion exchange resins from moderator and
primary heat transport purification systems. These resins are in temporary
storage, either at the plant of origin or at the designated radioactive waste
management sites.
18
TABLE 10. AVERAGE NORMALIZED AIRBORNE 14C RELEASESa
(GBq·GW(e)–1·a–1) FROM PWRs AND BWRs
1975–1979
1980–1984
1985–1989
PWR
222
345
120
BWR
518
330
450
a
Values are averages of all reported data and are taken from the 1982, 1988 and 1993
United Nations Scientific Committee on the Effects of Atomic Radiation (UNSCEAR)
reports [16–18].
Spent ion exchange resin from water purification systems at HWRs (e.g.
CANDU reactors) is a unique waste, owing to its relatively high 14C content (as
compared with that produced by LWRs). Typical specific activities of 14C on
resins of moderator purification systems are in the range 1.7–7.9 TBq/m3.
Carbon-14 activities on resins of heat transport purification systems range from
0.01 to 0.2 TBq/m3. Considerable quantities of these resins will be accumulated
at the end of a reactor’s lifetime.
Figure 3 shows the flows of 14C in an HWR. Annual 14C releases from two
Canadian CANDU (600 MW(e)) stations from 1991 to 1998 are shown in
Table 13.
Heavy water circulation
Fuel
Cladding
Storage
HWR
Reactor components
Moderator and coolant
Decommissioning
Ion exchange
resins
Annulus gas
Cover gas
Reprocessing plant
Storage
Absorption from off-gas
Immobilization
Discharge
Disposal
Environment
FIG. 3. Simplified diagram of 14C flows in an HWR and at the fuel reprocessing stage.
19
TABLE 11. PRODUCTION OF 14C (GBq·GW(e)–1·a–1)
IN CANDU 600 MW(e) REACTORS
Moderator
27 000
Primary heat transport
380
Fuel
960
Annulus gas (with CO2)
38
TABLE 12. ESTIMATED 14C DISTRIBUTION FOR CANDU 600 MW(e)
REACTORS
Total 14C production
(%)
Emitted 14C
(%)
14
Moderator
95.20
3.77
91.43
Primary heat transport
1.30
0.00
1.30
Fuel
3.37
0.00
0.00
Annulus gas
0.13
0.13
0.00
C on ion exchange
resins (%)
TABLE 13. ANNUAL AIRBORNE 14C RELEASES
(GBq/a) FROM TWO CANADIAN CANDU 600 MW(e)
REACTORS
20
Point Lepreau
Gentilly 2
1991
350
450
1992
720
—
1993
590
480
1994
200
2920
1995
142
1650
1996
121
—
1997
150
500
1998
320
277
3.1.3.3. Magnox reactors and advanced gas cooled reactors
Magnox reactors and AGRs have been the basis of the British nuclear
programme. Arisings of 14C in these reactors are dominated by the contribution
of the graphite moderator due to 13C(n,g)14C and 14N(n,p)14C reactions; for
example, approximately 1.85 TBq (50 Ci) of 14C is generated per year in a
250 MW(e) gas cooled graphite moderated reactor: about 60% is produced by
interaction with nitrogen impurities in the graphite and 40% by interaction
with 13C contained in the graphite pile [11].
Most of the graphite moderator remains in the reactor core in both
Magnox reactors and AGRs and will constitute a solid waste arising at the
decommissioning stage. However, the graphite sleeves of AGRs are removed
with the fuel and pass to the reprocessing plant, where they become a separate
solid waste arising. Bush et al. [12] calculated 14C production rates for Magnox
reactors and AGRs; the results are shown in Table 14.
An example of the total radionuclide inventory in a graphite moderator is
given in Table 15 [11].
The graphite becomes corroded during reactor operation by a radiolytic
reaction in the presence of the CO2 coolant. The rate of the reaction depends
on the irradiation level, the coolant composition and the pore structure of the
graphite. The reaction can be inhibited by adjustment of the coolant composition, but is still significant.
A large portion of the 14C gaseous releases from gas cooled reactors
comes from the purification of the CO2 circuits used to cool the reactor and
from isotopic exchange between the moderator and the CO2 circuit.
The estimated quantity of 14C released from the graphite moderator into
the coolant for Magnox reactors and for AGRs is provided in Table 16 [12].
Carbon-14 production within the coolant depends on the nitrogen
impurity level present. A level of 100 ppmv was used in Table 16, and the 14N
activation reaction accounts for 65% of the total 14C production in the coolant.
The total 14C inventory in the coolant is discharged into the atmosphere,
either by leakage or by a periodic routine purge. The remaining 14C, in the fuel
and its cladding, will pass to the reprocessing plant. Figures 4 and 5 show the
flows of 14C in Magnox reactors and AGR reactors, respectively. The fraction
released from the reactor is about 3% for Magnox reactors and about 6% for
AGRs of the total production of 14C in the reactors [15]. The emission level of
14
C from the RBMK type reactor in the Russian Federation has been estimated
to be 1000 ± 300 GBq·GW(e)–1·a–1 [19]. In France the activity of 14C released
from the Chinon A2 reactor (a 200 MW(e) reactor) is reported to be about 370
GBq/a, which is the same as from Saint Laurent A (a 450 MW(e) reactor) [11].
21
TABLE 14. CALCULATED 14C PRODUCTION RATES (GBq·GW(e)–1·a–1)
IN MAGNOX REACTORS AND AGRs
Fuel
Cladding
Coolant
Moderator
Total
Magnox
3 740
295
405
9 550
13 990
AGR
295
1 810
220
5 700
8 025
TABLE 15. TYPICAL RADIONUCLIDE INVENTORY (TBq) IN A
GRAPHITE MODERATOR AFTER 10 YEARS OF DECAY
Half-life (a)
Tritium
14
36
12.33
C
Cl
Marcoule G2 reactor Chinon A2 reactor
(1300 t)
(1700 t)
200
400
5 730
28
37
300 000
0.9
0.4
55
2.63
3
20
60
5.33
9
13
92
18
23
16
0.5
0.7
30
0.6
0.15
8.6/5
0.9
0.07
Fe
Co
63
Ni
93m
Nb
137
Cs
154
155
Eu/ Eu
TABLE 16. CARBON-14 ARISINGS (GBq·GW(e)–1·a–1) IN
THE COOLANT OF MAGNOX REACTORS AND AGRs
22
Magnox
AGR
Total production in graphite
9550
5700
Of which released to coolant
700
330
Production in coolant
405
220
Total present in coolant
1105
550
3.1.3.4. High temperature gas cooled reactors
High temperature gas cooled reactors (HTGRs) were the subject of
research and development work in Germany (in the former Federal Republic
of Germany) and in the USA. They present special problems in the field of the
management of waste containing 14C for the following reasons:
(a)
(b)
Releases of 14C to the environment from the HTGR fuel cycle per unit of
electricity generated will be much greater than for other types of reactor.
Owing to the structure of the fuel elements, much, or all, of the graphite
moderator will have to be removed as the first step in fuel reprocessing.
Most of the 14C will therefore be released, together with a large amount of
CO2 from the combustion of the graphite moderator, at the reprocessing
plant.
The 14C production rate in an HTGR reactor is highly dependent on the
nitrogen level in the graphite moderator. In general, HTGRs could give rise to
3700–7400 GBq·GW(e)–1·a–1 of 14C, eventually all of which will pass to the
reprocessing plant.
Gas circulation
Fuel
Magnox reactor
Cladding
Moderator (graphite)
Reprocessing plant
Decommissioning
Absorption from off-gas
Immobilization
Discharge
Disposal
Coolant
Purge and leakage
Environment
FIG. 4. Simplified diagram of 14C flows in a Magnox reactor and at the fuel reprocessing
stage.
23
3.1.3.5. Fast breeder reactors
There is no moderator in a fast breeder reactor (FBR), although the
coolant and other core components do have some moderating effect. The
coolant is liquid sodium. The significant sources of 14C would therefore be
oxygen in the fuel and nitrogen impurities in the fuel and cladding. Nitrogen
levels in the coolant will be very low, and releases of 14C from reactors will not
be significant. Essentially all the 14C inventory will pass to reprocessing plants
in the fuel elements.
Table 17 shows the typical ranges of 14C production rates in the fuel and
cladding of FBRs, based on data published in Ref. [12].
3.1.3.6. Summary of waste containing 14C produced by reactor operation
It can be summarized, based on the discussion in this section, that the
predominant quantity of 14C from nuclear power reactors occurs in the
following waste streams:
(a)
Reactor off-gases: the 14C is mainly produced in the moderator and
coolant during reactor operation. The chemical form of 14C is predominantly CO2 for HWRs, BWRs, AGRs and Magnox reactors, but predominantly hydrocarbons for PWRs.
Air circulation
Fuel
AGR
Cladding
Moderator (graphite)
Reprocessing plant
Decommissioning
Absorption from off-gas
Immobilization
Discharge
Coolant
Purge and leakage
Disposal
Environment
FIG. 5. Simplified diagram of 14C flows in an AGR and at the fuel reprocessing stage.
24
TABLE 17. CALCULATED 14C PRODUCTION
RATES (GBq·GW(e)–1·a–1) IN FBRs
Fuel
200–230
Cladding
295–475
Total
495–705
TABLE 18. ARISINGS AND RELEASES OF 14C FROM VARIOUS
TYPES OF REACTOR
Installed
capacity
(MW(e))
Gaseous waste Liquid effluent
(GBq/a)
(GBq/a)
Solid waste
(decommissioning)
(GBq/a)
LWR–PWR
1000
129.5
1.3
647.5
LWR–BWR
1000
259.0
1.3
1165.5
HWR
600
3108
Small
703a
GCR–MGR
480
373.7
Small
2982.2
GCR–AGR
660
255.3
Small
2479
GCR–HTGR
600
14.8
Small
Small
FBR
1250
0.65
Small
Small
a
Value taken from Dubourg [11].
(b)
Solid waste: this includes the graphite moderator in AGRs and Magnox
reactors, ion exchange resins in other types of reactor and structural
materials in the reactor core of all types of reactor.
In general, a small quantity of 14C is released into liquid effluents from
nuclear reactors. Estimated arisings and releases of 14C for various types of
reactor are summarized in Table 18 [12].
3.1.4.
Release during spent fuel reprocessing
In reprocessing plants a significant part (e.g. typically 40–50% for LWR
fuel) of 14C is released from spent fuel into the off-gas at the dissolution stage,
so measures should be taken for off-gas cleaning. The remaining 14C is retained
with insoluble cladding, which is further managed as solid waste.
The 14C behaviour and the possible pathways to the environment in a
spent fuel reprocessing plant are briefly discussed below, based on metal clad
25
LWR fuels. A typical flow diagram of an LWR fuel reprocessing plant is shown
in Fig. 6.
Spent fuel is placed upon arrival at a reprocessing plant in the fuel
receiving and storage pool. In general, no significant quantity of the 14C
arriving at the reprocessing facility is expected to escape to the fuel storage
pool, even in the case of defective fuels.
The first step in reprocessing metal clad fuels is shearing the fuel rods.
Fuel elements are mechanically transferred from the storage pool to the
remote process cell, in which shearing (chopping) occurs. Any 14C present in
the gas space inside the rods will be released into the shear cell off-gases at this
stage. Studies have shown that only a very small fraction of the total 14C
inventory in the fuel is released in this way [12].
Spent fuel
100% 3H and 14C
Nitrogen oxide
removal
t
t
Storage (cooling)
Off-gas
<1% 3H
40–45% 14C
t
Shearing
Dissolution
Discharge to
atmosphere
0.2–2.0% 14C
Iodine and
carbon dioxide
absorption
Carbonate
precipitation
Residual
carbon dioxide
removal
To storage or
cementation
Nitric acid
t
t
Cladding
60% 3H
55–60% 14C
Uranium and plutonium
fission products
40–45% 3H
40–45% 14C
Iodine
immobilization
Ba/Ca carbonates
38–45% 14C
Uranium
purification
cycle
Ha raffinate
40–45% 3H
Evaporation
Bottoms (HLW)
2–4% 3H
Storage
t
t
First extraction cycle,
uranium and
plutonium extraction
First extraction cycle,
scrubbing
Scrub
1–5% 3H
First extraction cycle,
uranium and
plutonium separation
Uranium product
~0.01% 3H
Plutonium
purification cycle
Various sources of ILW
Condensate
36–43% 3H
Purification cycle raffinates
~0.1% 3H
Evaporation
Condensate
40–45% 3H
Fractionation
Bottoms (ILW)
Bottoms
(nitric acid)
<5% 3H
Condensate
2–4% 3H
Solidification
Disposal
Plutonium
~0.1% 3H
Condensate
40% 3H
Discharge
(to sea or pond)
or disposal
To fuel
dissolution
FIG. 6. Simplified flow diagram for a typical plant for LWR fuel reprocessing (the tritium
and 14C contents in the process products are expressed as a percentage of their amounts in
spent fuel).
26
The fuel sheared into segments containing hulls and oxide pellets is then
transferred to the dissolver, where the pellets react with nitric acid. Studies
have indicated that under dissolution conditions essentially all the 14C from the
dissolver liquor is evolved into the gas phase. This is an important result,
because any 14C compounds retained in the dissolver solution might be
extracted in the subsequent solvent extraction process, and might then build up
in the recycling solvent. Under the oxidizing conditions in the dissolution step,
carbon is almost completely oxidized to CO2, which is liberated into the
dissolver off-gases. The 14C content in dissolver liquor is insignificant compared
with the amounts in dissolver off-gases and in the fuel cladding.
Depending on the design and efficiency of the gas cleaning system, 14C is
either separated from the dissolver off-gases and converted to solids for
disposal (as shown in Fig. 6) or discharged into the atmosphere.
Studies have shown that essentially all the 14C in dissolver off-gases is
present in the dioxide form [12]. In particular, the predominance of CO2 as the
14
C species present in the off-gases of a reprocessing plant is confirmed by
measurements of emissions at the Karlsruhe plant in Germany. Emissions of
_
14
C as CO2 when reprocessing PWR and BWR fuels were 459 GBq·GW(e) 1·a–1
and 640 GBq·GW(e)–1·a–1, respectively, being in good agreement with the
expected production in these fuels [12].
Studies of emissions during the reprocessing of fuel from the experimental HWR at Karlsruhe gave similar results to those for LWR fuel. Dubourg
[11] has estimated that the 14C activity released from reprocessing the fuels
from the Chinon A2 reactor (with a 200 MW(e) capacity) is about 555 GBq/a
and that from Saint Laurent A (with a 450 MW(e) capacity) is 1110 GBq/a. No
data are available on the behaviour of 14C present in FBR fuel (U/Pu oxide),
but similar behaviour to that in LWR fuel can probably be expected [13]. Most
of the 14C from HTGRs will be released during the reprocessing of spent fuel.
Table 19 [14] shows estimates of 14C emissions and inventories in solid
waste for various reactor systems and reprocessing plants.
It can be concluded from the above information that most 14C will occur
in the following waste streams at fuel reprocessing plants:
(a)
(b)
(c)
Off-gases released during fuel dissolution: these will generally contain the
14
C originally generated and collected in the fuel. This could either be
released as CO2 gas, scrubbed out and discharged as a liquid, or converted
to a solid.
Burner off-gases at HTGR fuel reprocessing plants: these will contain the
significant 14C content of the graphite moderator (integral with the fuel
elements) in a gas stream rich in CO2.
Solid waste: the fuel cladding.
27
TABLE 19. ESTIMATED 14C EMISSIONS AND INVENTORIES IN
SOLID WASTE FOR VARIOUS REACTOR SYSTEMS
Reactor
Power
(MW(e))
PWR
900
Reprocessing plant
Estimated Airborne
14
14
Solid waste
C
C
(GBq/a)
production emission
(GBq/a)
(GBq/a)
1850
148
666
Airborne
14
Solid waste
C
(GBq/a)
emission
(GBq/a)
518
518
PWR
1300
2590
185
925
740
740
BWR
1000
3515
296
1295
962
962
Graphite
reactora
200
2849
370
1850
555
74
Graphite
reactorb
450
4255
370
2664
1110
266
AGR
600
4070
296
2664
185
925
HWR
600
5550
4070
—
814
703
Note: The calculations were based on 7000 operating hours per year.
a
Chinon nuclear power plant (1700 t graphite moderator) in France.
b
Saint Laurent A2 nuclear power plant (2440 t graphite moderator) in France.
Table 20 summarizes the expected specific quantities of 14C in gaseous
and solid waste for the types of reactor considered and for corresponding
reprocessing plants [12].
3.2. TRITIUM PRODUCTION AND RELEASE
Tritium is produced by nuclear reactions that occur naturally in the
environment, in nuclear weapon testing and in nuclear reactors. It is continuously generated by the interaction of high energy cosmic rays with oxygen and
nitrogen atoms in the upper atmosphere. These processes produce most of the
world’s natural tritium. Tritium converts into water and reaches the Earth’s
surface as rain. An estimated production rate of 1.48 × 108 GBq/a results in a
world steady state natural inventory of 2.59 × 109 GBq [20].
Atmospheric nuclear test explosions from 1945 to 1975 added about 2.96
11
× 10 GBq of tritium to the environment, much of which has since decayed.
However, about 1.85 × 1010 GBq (5 × 108 Ci) remains in the environment,
28
TABLE 20.
EXPECTED NORMALIZED 14C DISTRIBUTION
–1 –1
(GBq·GW(e) ·a ) IN WASTE AT NUCLEAR FACILITIES
Gaseous wastea
LWR–PWR
Solid waste
Total produced
Reactor
Reprocessing
plant
Reactorb
Reprocessing
plant
2 590
185
740
925
740
LWR–BWR
3 515
370
740
1 665
740
HWR
10 175
7 400
1 480
NR
1 295
GCR–MGR
14 060
1 110
3 700
8 880
370
CGR–AGR
8 140
555
370
5 365
1 850
CGR–
HTGR
5 587
37
5 550
Small
Small
FBR
925
185
NR
740
0.74
a
This waste could be discharged as gases, transferred to liquid effluents or converted to
solids for disposal.
b
Decommissioning waste.
NR: no data have been reported.
mostly diluted in the oceans. Underground nuclear tests appear to add little
tritium into the atmosphere [20].
3.2.1.
Production and release in nuclear power reactors
Production processes, which are described in detail in Refs [21–23], and
mechanisms for the generation of tritium at nuclear reactor facilities have not
fundamentally changed in the past decade. Tritium is produced in reactors by
neutron activation of 2H, 3He, 6Li and 10B, and is thus formed in the following
reactor components [24]:
(a)
(b)
(c)
(d)
(e)
(f)
(g)
The water coolant–moderator (LWRs);
The heavy water coolant and moderator (HWRs);
The helium coolant (HTRs);
The boron control rods (many types of reactor);
The dissolved boric acid in the moderator (PWRs, HTRs);
The graphite moderator (lithium impurity: gas cooled reactors (GCRs),
HTGRs);
The UO2/PuO2 core and UO2 breeder fuel (lithium and boron impurities:
FBRs).
29
TABLE 21.
ESTIMATED TRITIUM PRODUCTION
(GBq·GW(e)–1·a–1) IN VARIOUS TYPES OF REACTOR
RATES
Fuel
Coolant
Moderator
Total
LWR–PWR
5.18 × 105
3.70 × 104
NA
5.55 × 105
LWR–BWR
5.18 × 105
Low
NA
5
5.18 × 105
5.18 × 10
5.42 × 107
Low
(0–1.85) × 105
(5.18–7.03) × 105
5.18 × 105
1.85 × 105
(0.18–7.40) × 104
(5.2–5.9) × 105
7.40 × 105
7.40 × 104
NA
8.14 × 105
HWR
5.18 × 10
GCR
5.18 × 105
GCR–HTGR
FBR
1.85 × 10
6
7
NA: data not available.
The estimated tritium production rates in various types of reactor are
reported in Ref. [25] and summarized in Table 21.
Most of the fission product tritium produced in fuel rods is usually
retained within the fuel and is not released to the environment at the reactor
site; it is released only during fuel reprocessing, if fuel reprocessing is carried
out.
Tritium in reactor core materials may remain where it is formed and be
accumulated until the reactor is decommissioned. The quantities of tritium in
various reactor components were estimated in Ref. [25] and are shown in
Table 22.
The most important movement of tritium in the reactor is diffusion of a
portion of tritium from the fuel either into the zircaloy cladding, where it is
trapped (most water cooled reactors), or through stainless steel cladding into
the coolant (AGRs, FBRs). Only a small quantity of tritium escapes from
Magnox fuel elements. The portion of tritium trapped in zircaloy cladding is
typically 15–60%, while about 30% passes into the GCR coolant and >95%
into the FBR (sodium) coolant [25]. A further transfer of tritium from the
graphite moderator of GCRs to the coolant may occur due to corrosion of the
moderator.
The activity produced in the coolant is partly or entirely released in the
effluent streams, depending on the waste management practices at the plant.
Releases to the environment are mainly in the form of HTO in reactors that use
water as the primary coolant. The typical discharge rates of tritium from
various types of reactor are summarized in Table 23 [25].
30
TABLE 22. ACCUMULATION OF TRITIUM (GBq/kW(e))
IN VARIOUS REACTOR COMPONENTS
Heavy water coolant–moderator (HWRs)
7.40 × 108
Graphite moderator (GCRs, HTRs)
3.70 × 104
Boron control rods
1.85 × 106
Cold traps for sodium purification (FBRs)
1.11 × 107 a
a
Transfer from fuel.
3.2.1.1. Light water reactors
The main source of tritium in LWRs is ternary fission, which produces
(5.55–7.40) × 105 GBq·GW(e)–1·a–1 [23]. This tritium is partly transferred with
the fuel to the reprocessing plant. The amount of tritium produced through
neutron reactions in the coolant–moderator is small, less than 3.70 × 104
GBq·GW(e)–1·a–1 [26]. The formation of tritium by activation reactions in
PWRs is considered to be mainly from boron in the coolant water, which is
used for reactivity control. In BWRs the contribution of tritium by activation
reactions is mainly from boron in the control rods. The tritiated effluents from
light water (the coolant–moderator of LWRs) can be released directly to the
environment without additional processing, owing to the very low concentrations of tritium.
The average normalized tritium releases from PWRs and BWRs between
1975 and 1989 are summarized in Table 24. It can be noticed that tritium
releases from PWRs gradually decreased by almost a factor of three between
1975–1979 and 1985–1989. This decrease is the result of better maintenance and
management of the reactor component systems.
Tritium activities discharged from BWRs to the environment are lower
than those of PWRs because less tritium is produced in or diffuses into the
primary coolant.
3.2.1.2. Heavy water reactors
The amount of tritium generated in fuel by ternary fission is approximately the same in HWRs as in LWRs. However, a relatively large amount of
tritium is produced in the D2O coolant and moderator (approximately 8.9 ×
107 GBq·GW(e)–1·a–1, primarily in the operation of the moderator system [26].
In a CANDU 600 MW(e) HWR most of the tritium is formed in the
thermal neutron capture reaction, 2H(n,g)3H. The reaction occurs in the heavy
31
_
TABLE 23. TYPICAL TRITIUM DISCHARGE RATES (GBq·GW(e) 1·a–1)
FROM VARIOUS TYPES OF REACTOR
Gaseous effluent
Liquid effluent
PWR (zircaloy cladding)
3.70 × 103
2.59 × 104
BWR
1.85 × 103
3.70 × 103
HWR
7.40 × 10
5
1.85 × 105
GCR
7.40 × 103
1.11 × 104
water of both the moderator and coolant systems. The mass of heavy water in
the high flux zones of the fuel channels is small compared with the mass of
moderator heavy water in the reactor core. In addition, the average thermal
neutron flux inside the fuel channels is slightly lower than that in the
moderator. As a result, the rate of tritium formation is much higher in the
moderator than in the coolant system. This production of tritium results in a
steady rise of tritium concentration in the heavy water inside the moderator;
for example, the calculated concentration of tritium in the moderator heavy
water is 3640 GBq/kg after 40 years, while in the coolant heavy water it is only
81 GBq/kg, for the same operation time, based on a capacity factor of 90% [27].
Leakage of water from the moderator and coolant systems gives rise to
tritium release to the reactor building. In general, most of the heavy water
leakage is recovered, for economic and radiological reasons, so that tritium
emissions to the environment by way of airborne and water emissions are kept
very low. A portion of the heavy water from the coolant and moderator systems
of HWRs may also be withdrawn for detritiation before it is returned to the
reactor, and some tritium may escape during this process; for example, the
average annual airborne tritium emission from a CANDU 600 MW(e) reactor
(Point Lepreau nuclear power plant, Canada, 1984–1994) was 2.68 × 105 GBq/a.
The average annual aqueous emission was 1.85 × 105 GBq/a [27].
3.2.1.3. Other types of reactor
In GCRs the contribution of tritium is from lithium impurities in the
graphite moderator and from the presence of water vapour in the core. In
Magnox reactors CO2 in the coolant circuit is continually dried to remove
water vapour, and hence tritium produced in the core finds its way primarily
into the liquid effluent removed by the dryers.
The five year average normalized airborne tritium releases from various
types of reactor are compared in Table 25; as expected, HWRs release more
32
TABLE 24.
AVERAGE NORMALIZED AIRBORNE TRITIUM
_
_
RELEASESa (GBq·MW(e) 1·a 1) FROM PWRs AND BWRs
1975–1979
1980–1984
1985–1989
PWR
7.8
5.9
2.8
BWR
3.4
3.4
2.5
a
Values are averages of all reported data taken from the 1982, 1988 and 1993
UNSCEAR reports [16–18].
airborne tritium than the other types of reactor. It should be noted that the use
of large amounts of heavy water as a moderator and coolant is one of the major
design features of HWRs for improved neutron efficiency, which inherently
causes a higher production and release of tritium.
3.2.2.
Release during spent fuel reprocessing
The amount of tritium in the fuel that passes to a reprocessing plant
depends on the type of reactor and on the fuel cladding material. Most of the
tritium released from the fuel during dissolution appears in the liquid waste
streams, while some is found in the dissolver off-gas stream and a portion is
immobilized as a solid zirconium compound in the cladding. Fuel with stainless
steel cladding may lose up to 50% of the tritium produced in the fuel because of
diffusion and permeation through the cladding. In AGRs and FBRs with
stainless steel clad fuel, only small amounts of tritium are expected to be passed
to the reprocessing plant. Zircaloy clad fuel, the fuel of most LWRs, retains
essentially all the tritium through the formation of zirconium tritide. For an
LWR fuel reprocessing plant, with a capacity of 1400 t/a of heavy metal serving
50 GW(e) of electric production capacity, the tritium release rate was estimated
to be (1.85–3.70) × 107 GBq/a (0.5–1.0 mCi/a) [26].
In an LWR fuel reprocessing plant using the Purex process (a typical flow
diagram of an LWR fuel reprocessing plant is shown in Fig. 6), after chopping
the fuel elements and fuel dissolution in nitric acid, about 60% of the total
amount of tritium is retained in zircaloy cladding [28]. The bulk of the
remainder is in the form of HTO and nitric acid, which is distributed through
the various streams in the process. A simplified distribution of tritium in a
conventional fuel reprocessing plant is shown in Table 26 [26].
As fresh water and nitric acid are added to the process, the tritium
becomes progressively more diluted, so large volumes of tritiated aqueous
effluents, of the order of 100 m3/t of heavy metal, can be produced. In reproc-
33
TABLE 25.
AVERAGE NORMALIZED AIRBORNE TRITIUM
_
_
RELEASESa (GBq·MW(e) 1·a 1) FROM VARIOUS TYPES OF REACTOR
1975–1979
1980–1984
1985–1989
PWR
7.8
5.9
2.8
BWR
3.4
3.4
2.5
HWR
540
GCR
—
670
5.4
480
9.02
FBR
—
—
96
Light water cooled
graphite moderated
reactor
—
—
26
a
Values are averages of all reported data taken from the 1982, 1988 and 1993
UNSCEAR reports [16–18].
essing plants at coastal sites these large volume effluents are discharged
directly into the sea, in which there is an enormous additional dilution effect.
Solar evaporation of HTO is used in India, while long term storage in open
ponds is used in the Russian Federation.
Table 27 provides typical annual amounts of tritium (per a unit of electric
power) in waste streams and process products from an LWR reprocessing
plant.
Concepts for concentrating tritium into smaller volumes of water have
been developed for locations where access to the sea is limited or discharge is
otherwise prohibited. Effluent volumes can be reduced to the order of 1 m3/t of
heavy metal by means of partial recycling of the water and nitric acid, along
with segregation of tritium in the first extraction cycle of the Purex process.
Use of the voloxidation process (see Section 5) as an alternative head end
process (dissolution of fuel in nitric acid) could further reduce the volume of
tritiated wastewater to the range of 2 to 20 L/t of heavy metal, with a tritium
concentration in the range 370–3700 GBq/L (10–100 Ci/L) [26]. This highly
tritiated water can be then converted into the less toxic HT form by electrolysis
and discharged into the atmosphere. However, for reasons of cost and owing to
some technological drawbacks, the voloxidation process has not been
employed in operating reprocessing plants.
34
TABLE 26. SIMPLIFIED DISTRIBUTION SCHEME OF TRITIUM IN A
CONVENTIONAL FUEL REPROCESSING PLANT
Operation
Process stream
Fraction of tritium in the
process stream (per cent of
inventory)
Chopping
Fuel element
Fuel element pieces
Off-gas
100a
>99.99
<0.001
Dissolution
Aqueous phase
Cladding
Off-gas
~40
~60b
<0.5
Extraction
Aqueous phase
Organic phase
>35
<5
Re-extraction
Aqueous phase
Organic phase
Plutonium product
Uranium product
<4.9
~0.01
~0.1
~0.01
a
b
Corresponding to 350–700 Ci/t at 30 GW(e)·d·t–1.
Depends on local fuel rod temperatures.
TABLE 27. TYPICAL TRITIUM DISTRIBUTION
_
_
(GBq·GW(e) 1·a 1) AMONG LWR REPROCESSING
PLANT WASTE STREAMS AND PROCESS
PRODUCTS
Solid waste (zircaloy cladding)
3.3 × 105
Liquid waste (aqueous)
~2.2 × 105
Liquid waste (spent solvent)
<55
Gaseous waste
2.8 × 103
Plutonium product
5.5 × 102
Uranium product
55
Total
5.5 × 105
35
3.3. CATEGORIES OF WASTE CONTAINING 14C AND TRITIUM
3.3.1.
Reactor operation and fuel cycle waste
Most of the human-made 14C and tritium will continue to be produced in
nuclear power reactors. The fate of these contaminants depends highly on the
reactor type and on its design and operation. Typical production rates for
various types of reactor are shown in Table 5. Fractions of these radioisotopes
will go into different waste streams produced in the nuclear fuel cycle. In
general, the following waste streams that contain 14C and tritium will be
generated routinely by reactor operations:
(a)
(b)
(c)
(d)
(e)
(f)
(g)
(h)
(i)
Reactor coolant water (liquid);
Spent ion exchange resins (solid);
Spent filter cartridges (solid);
Radioactive refuse (solid);
Reactor off-gases (gas);
Condensates from off-gas systems (liquid);
Coolant (CO2) of Magnox reactors and AGRs (gas);
Highly irradiated steel structures and components (solid, decommissioning waste);
Graphite moderators, mainly from graphite reactors (solid, decommissioning waste).
There are a number of waste streams containing 14C and tritium
associated with the reprocessing of spent nuclear fuel, which include:
(i) Fuel cooling storage basin water (liquid);
(ii) Spent fuel washing solutions and condensates from head end fuel
treatments (liquid);
(iii) Dissolver off-gases from the gas cleaning system (gas);
(iv) Off-gas scrubber solutions, which are partially recycled (liquid);
(v) Condensates from off-gas systems (liquid);
(vi) Highly active raffinate from the first extraction cycle of uranium and
plutonium decontamination after it is concentrated by evaporation
(liquid);
(vii) Recovered nitric acid that is recycled for fuel dissolution (liquid);
(viii) Recovered water (condensate) from the nitric acid regeneration process
(liquid);
(ix) Refuse from the plant’s operation (solid).
36
3.3.2. Other waste
Waste containing 14C and tritium is also generated by radioisotope
producers, hospitals, medical schools, universities and colleges. Such waste can
be in the form of refuse, liquid scintillation vials, absorbed aqueous and organic
liquids, discard sealed sources and biological waste.
The inventories of 14C and tritium, and volumes of the waste generated, in
radioisotope production for industrial uses and medical and research activities
are much smaller than those generated by reactor and fuel reprocessing
operations. Treatment of this waste by the technologies discussed in the
following sections may not be required.
Waste generated in research activities such as accelerator operations will
contain tritium and/or 14C. However, the total volume of this waste is likely to
be small.
There are a number of historical waste streams (e.g. equipment related to
nuclear testing programmes) around the world that may contain 14C and
tritium. In particular, in the past a large quantity of tritium was used in the
production of luminous materials, electrostatic neutralizers, etc. While 14C and
tritium are not necessarily the most significant radionuclides in this type of
waste, they may need to be considered in its treatment and disposal.
4. REDUCTION OF PRODUCTION AND RELEASE
4.1. REDUCTION OF 14C PRODUCTION AND RELEASE
It has been suggested that some 14C production sources may be controlled
by limiting the amounts of parent substances subject to irradiation. Such
control would also reduce releases of 14C from nuclear reactors [13].
In general, two neutron activation reactions dominate the production of
14
C: 17O(n,a)14C and 14N(n,p)14C, which, respectively, have a cross-section of
0.24 barn and 1.82 barn. Most of the 14C produced in fuel will pass to the
reprocessing plant and will be released with the dissolver off-gases. It is
therefore useful to consider the possibility of reducing the initial nitrogen and
17
O contents in the fuel.
_
_
Calculations indicate that approximately 0.15 TBq·GW(e) 1·a 1 of 14C is
produced by the activation of 17O in LWR fuel. This represents an irreducible
minimum, since the amount of oxygen present in the fuel is somewhat
37
inflexible to adjustment. However, the nitrogen impurities in the fuel may be
controlled during fuel fabrication. Bush [29] indicated that a modest reduction
(by at most a factor of about five) in 14C production could be achieved by a
reduction in the level of nitrogen impurity in the fuel; for example, at a level of
_
_
25 ppm of nitrogen in the fuel, nitrogen contributes 0.55–0.74 TBq·GW(e) 1·a 1
of 14C. Reduction of this level to 5 ppm (a level frequently found in some fuels)
gives a contribution from nitrogen about equal to that from 17O.
The materials of core structures contain certain amounts of nitrogen (e.g.
the nitrogen contents in stainless steel range from 0.04% to 0.08%). However,
nitrogen levels in many reactor materials are not known with certainty, and it is
therefore difficult to estimate what reduction might be possible. The costs of
decreasing the nitrogen levels in reactor core materials should be balanced
against the benefit of a better mechanical property and corrosion resistance
and the reduction in collective doses to humans thereby achieved. The total
cost of reducing nitrogen levels plus that of the remaining dose detriment
should be minimized in accordance with the ALARA (as low as reasonable
achievable) principle.
Gaseous 14C is produced in LWRs mainly by activation of nitrogen
entering the primary coolant, and is released in gaseous effluents. The average
PWR release rate is about 185–370 GBq of gaseous 14C per year. In order to
minimize the presence of nitrogen in the primary coolant, the cover gas of
water storage tanks has been switched from nitrogen to argon, so as to prevent
the dissolution of nitrogen gas in the primary coolant. Recently, operating
experience of HWRs in the Republic of Korea has indicated that controlling
the nitrogen level in the moderator cover gas may lead to lower 14C emissions.
Moreover, the pH control of the primary coolant in Western PWRs is
carried out by depleted 6Li, lithium hydroxide, instead of hydrazine
(NH2_NH2), to prevent the formation of 14C. Better control of the chemicals
used in the operation of reactors is one means of reducing 14C production.
Air ingress, especially in the moderator and primary coolant systems, has
been found to be one of the sources of 14C production during reactor operation.
Improving the system design to minimize the potential for air ingress will
therefore reduce 14C production and release.
It is interesting to note that the old style CANDU reactors (using
nitrogen as the annular gas), WWER reactors and RBMK reactors released
large quantities of 14C in gaseous effluents, owing to the extensive use of
nitrogen blanketing inside the reactor. The current CANDU reactors, using
CO2 as the annular gas, have shown a significant reduction in 14C production in
the annular gas system. Avoiding the use of chemicals that could produce 14C
during reactor operations is a viable approach to reducing 14C production.
38
In general, HWRs produce relatively large quantities of 14C as compared
with LWRs; this is because:
(a)
(b)
The quantities of water exposed to neutron irradiation are much larger
than in LWRs;
The heavy water contains more 17O (about 55% above the natural level),
since enrichment of 17O occurs during the production of heavy water.
Reduction of 17O to less than the normal level of 0.038% is likely to be
uneconomic, but restoring the oxygen in heavy water to the natural concentration could be a practicable proposition. Again, the ALARA principle
should be applied to balance the costs of reducing the 17O levels in the heavy
water against the benefit of the reduction in collective doses to humans.
One of the pathways for 14C emissions from HWRs is venting and purging
the moderator cover gas and the heat transport system. The frequency of
venting and purging can be reduced by improving reactor operation practice,
which will reduce the release of 14C to the environment.
4.2. REDUCTION OF TRITIUM PRODUCTION AND RELEASE
Tritium is produced in all nuclear reactors as a by-product of ternary
fission, which is the largest source of tritium production in power reactors.
Tritium is also formed as a product of other reactions, such as reactions with 2H,
3
He, 6Li and 10B.
As the largest source of tritium production is within the fuel itself, which
typically comprises uranium, plutonium or thorium, it is desirable to keep the
tritium produced in close proximity to this source. One way to lower the tritium
released into the reactor coolant, and hence the amount of tritium that may be
discharged to the environment, is to provide a means within each fuel rod to
store the tritium produced.
It has been found that tritium diffuses through stainless steel in a reactor
environment at a high rate, the rate being significantly higher than tritium
diffusion through zirconium alloys. Tritium also reacts with the zirconium alloy
cladding to form a hydride, lessening the release of tritium to the reactor
coolant. An apparatus [30] has been developed to remove and store tritium
from a gaseous medium; this apparatus can be incorporated in a fuel rod to
retain tritium and to minimize the release of tritium to the reactor coolant. The
apparatus consists of an internal core of zirconium or zirconium alloy, which
retains tritium, and an adherent nickel outer layer, which acts as a protective
and selective window for the passage of the tritium.
39
Increased production of tritium in LWRs has been associated with the
presence of deposits known as crud (the word originated as an acronym for
Chalk River unidentified deposits). It was initially observed as a black deposit
on the first naval nuclear fuel test rod tested in the NRX reactor at Chalk
River, Canada. Crud has now become a standard nuclear industry term that
refers to minute, solid, corrosion products that become highly radioactive and
plate out on reactor rods, components and piping. The actual composition of
crud is unique to the water chemistry and fuel failure history of each reactor
[31].
Crud deposits consist mainly of iron oxides with high porosities. Uranium
from failed fuel elements (tramp fuel), boron and lithium are deposited or
trapped in the crud. These materials are concentrated on the surface of the fuel
rods, increasing the probability of neutron interactions, which result in an
increase in the production of tritium, fission products and activated corrosion
products.
In order to minimize the production of tritium and release into the
primary coolant of reactors, it is important to:
(a)
(b)
(c)
(d)
(e)
Improve fuel rod fabrication and the quality of the fuel rods, in order to
avoid fuel cladding failure.
Use zircaloy cladding instead of stainless steel cladding, where possible
(PWRs, BWRs).
Use enriched (99%) 7Li as lithium hydroxide for pH control in PWRs; it
should be noted that the predominant neutron activation reaction in
LWRs is 6Li(n,a)3H, with a cross-section of 942 barn.
Isolate as soon as possible defect fuel assemblies in leaktight canisters, in
order to avoid polluting spent fuel pools.
Control the coolant chemistry, in order to minimize the deposit of crud on
the fuel rods.
In order to prevent the spread of tritium contamination, it is advisable to
monitor nuclear sites with tritium measurement systems in several locations, to
use impermeable layers under the reactor units to prevent dispersal of tritium
in the aquifer, and to allow for the possibility of collecting contaminated water
in separate tanks.
40
5. TECHNOLOGIES TO CAPTURE 14C AND TRITIUM
FROM EMISSION STREAMS
One of the objectives of radioactive waste management is to minimize or
eliminate releases of contaminants to the environment and to reduce the final
waste volume for storage and/or disposal. In general, waste can be housed in
containers that provide containment for minimizing the release of contaminants (e.g. tritium and 14C), and volume reduction can be achieved through
processes such as compaction or incineration.
Since contaminants in liquid or gaseous waste are mobile and easily
migrate to the environment, methods are required to remove these contaminants from the waste streams and to convert the contaminants into a stable
waste form (e.g. a solid matrix) for storage and/or disposal. The following
sections review technologies for the removal of 14C and tritium from gaseous
and liquid waste streams.
5.1. REMOVAL OF 14C FROM GAS STREAMS
For physical and chemical reasons, CO2 is the carbon compound that can
most easily be separated from other gases. It is therefore prudent to oxidize
other 14C compounds into 14CO2 before its removal [32]. The treatment
processes discussed below focus on the removal of CO2 from gas streams.
The selection of a specific process for the capture and retention of CO2
from a gaseous stream depends on the volume of gas to be treated, the concentration of CO2 in the gas stream, the composition of the gas stream and the
desired final waste form. Some of the criteria used in the selection of removal
technology options include:
(a)
(b)
(c)
Cost considerations;
Simplicity of implementation;
Worker safety.
Various methods for 12CO2/14CO2 removal from gas streams have been
studied and have undergone different levels of development. Detailed reviews
are made in Refs [10, 12, 33], which examine proven and potential 12CO2/14CO2
removal processes in fuel reprocessing facilities, nuclear power plants and
industrial chemical separation plants that handle CO2. Methods for the
41
removal of 12CO2/14CO2 from gaseous effluents can be broadly divided into
four major categories:
(i) A single-step chemical reaction involving absorption in an alkaline earth
hydroxide slurry or a solid (e.g. an agitated slurry scrubber or a packed
bed column);
(ii) A two-step chemical reaction involving sodium hydroxide and lime slurry
(e.g. a double alkali process);
(iii) Physical absorption (e.g. absorption in water, absorption in methanol or
absorption in a fluorocarbon solvent);
(iv) Physical adsorption on an active surface (e.g. a molecular sieve process).
5.1.1.
Single-step chemical reaction involving absorption in an alkaline earth
hydroxide slurry or solid
Two processes in which CO2 is removed by absorption in an alkaline earth
hydroxide medium have been studied extensively and have advanced to a stage
ready for demonstration. One process is based on an agitated slurry scrubber
configuration and the other is based on a solid packed bed column design. Brief
descriptions of these two processes are given below.
5.1.1.1. Alkaline slurry scrubber
Essentially a complete removal of 12CO2/14CO2 from gas streams can be
accomplished by reaction with agitated alkaline slurries. The carbonate product
is very stable and satisfies many long term storage and/or disposal requirements. The feasibility of CO2 removal by contacting CO2 with Ca(OH)2 or
Ba(OH)2 slurries in an agitated gas–liquid contactor has been the subject of
several studies [34–37]. Agitated slurry scrubbers are often preferred to gas–
liquid contacting because of the high interfacial area available for mass transfer
(30 m2/m3) as compared with other processes, such as packed beds (0.5–3.0 m2/
m3) [10].
In general, a slurry scrubber consists of a variable-speed agitation system,
a gas distribution system and a stirring tank. The slurry concentrations range
from 10 wt% to 20 wt%, depending on the application.
Holladay [35] studied the removal of CO2 with lime slurry in a stirred
tank reactor (0.196 m internal diameter and 0.33 m high) and reported that the
rate of reaction was fast and that a virtually complete removal of CO2 could be
achieved. The decontamination factor was >100 in a single stage reactor for
CO2 concentrations ranging from 5% to 100% (pure CO2). The reaction rate
was constant up to a 90% utilization of the lime and then rapidly decreased, as
42
did the pH for the remainder of the reaction. Patch et al. [37] studied the
removal of CO2 from an air stream in a continuously sparged, agitated slurry
carbonation reactor (0.273 m internal diameter and 0.35 m high). Experimental
results indicated that the decontamination factor ranged from 3 to 1200 for
Ca(OH)2 slurries and from 60 to 2100 for Ba(OH)2 slurries. For both slurries
low values of decontamination factor were observed at higher gas flow rates
and lower impeller speeds.
Atomic Energy of Canada Limited (AECL) has recently developed a
prototype lime slurry scrubber for the removal of 14CO2 from the moderator
cover gas in the CANDU system. The scrubber was tested for more than 450 h
at various operating conditions using a specially designed helium recycle loop.
The results indicated that the slurry scrubber could remove low concentrations
of CO2 (0.04–1000 ppmv of 14CO2 + 12CO2) from gas streams with a high single
pass removal efficiency, high reagent utilization, relatively low system pressure
drop and phase entrainment.
5.1.1.2. Alkaline packed bed column
Absorption of CO2 in solid absorbents (e.g. alkaline earth hydroxides) is
another CO2 removal process. The absorbent is usually arranged in a fixed bed
configuration with the gas stream passing through the bed: the gas stream
enters the bed, CO2 is absorbed by the absorbent and the gas stream exits,
leaving CO2 behind.
A number of absorbents, belonging to the group I (alkali metal) and
group II (alkaline earth) hydroxides, can remove CO2 effectively from gas
streams; they include ascarite (NaOH on asbestos), LiOH·H2O, Ca(OH)2,
Ba(OH)2, soda lime (NaOH–Ca(OH)2 mixtures) and baralyme (Ca(OH)2–
Ba(OH)2 mixtures). However, some absorbents are not suitable for 14CO2
removal, owing to poor reactant utilization or a lack of a stable final product
for storage and disposal (e.g. ascarite, LiOH·H2O).
A number of feasibility studies for the removal of CO2 by Ca(OH)2
[38–42] and Ba(OH)2 [36, 43–45] in a fixed bed configuration have been
reported. The studies found that the efficiency of this process is closely related
to the amount of excess water in the influent gas.
Researchers at Ontario Hydro (now Ontario Power Generation)
developed a process for 14C removal based on the CO2–Ca(OH)2 reaction. The
process was designed for the removal of 14CO2 from the moderator cover gas of
CANDU reactors [40]. Over 1000 h of pilot scale testing were performed and
the test results showed that CO2 could be effectively removed by Ca(OH)2 at
ambient temperature, but a high utilization of absorbent could only be
achieved at relatively high humidity (80–100% relative humidity). The process
43
was further developed to remove 14C from the extremely dry nitrogen annulus
gas of CANDU nuclear power plants [46].
A demonstration unit based on the solid Ca(OH)2 fixed bed configuration has been built and successfully demonstrated for 12CO2/14CO2 removal
from the moderator cover gas in the Nuclear Power Demonstration reactor in
Canada [47]. The unit was operated from 27 April to 22 May 1987 for a total of
612 h. No CO2 was detected downstream of the unit, and a total of 234 g of
14
CO2 and 12CO2 was captured during the demonstration.
A subsequent demonstration of a modified solid Ca(OH)2 fixed bed
scrubber for 12CO2/14CO2 removal from the annulus gas system was carried out
in 1989 at Pickering generation station (unit 3) in Canada. During the demonstration period a total of 0.56 TBq of 14CO2 from the inventory, in addition to
0.17 TBq produced in seven days, was removed by the scrubber. The test results
indicated that no 14CO2 was detected downstream of the scrubber [47].
A process utilizing fixed bed canisters of Ba(OH)2·8H2O has been
developed at the Oak Ridge National Laboratory (ORNL), Tennessee, USA
[43–45]. The study at the ORNL was designed for the treatment of air based
(~350 ppmv CO2) gas streams, such as those potentially present in LWRs. The
CO2 removal systems (0.102 m internal diameter and 0.305 m internal
diameter) were tested over 18 000 h at isothermal conditions and over 2100 h at
near adiabatic conditions. The researchers at the ORNL concluded that the
Ba(OH)2·8H2O process is capable of high CO2 removal efficiencies and high
reactant utilization, and has acceptable operational characteristics at near
ambient conditions.
Recently, AECL has developed a prototype dry bed 14C removal system.
The system consists of layers of commercially available absorbent to capture
14
CO2. The 14C removal system was tested for more than 200 h at various
operating conditions. The test results showed that the dry scrubber could
remove low concentrations of CO2 (0.04–1000 ppmv of 14CO2 + 12CO2) from
gas streams with a high single pass removal efficiency, low system pressure drop
and phase entrainment.
5.1.2.
Two-step chemical reaction involving sodium hydroxide and lime slurry
Scrubbing is a popular commercial method for the removal of one or
more constituents from a gaseous stream and involves absorbing into a liquid
stream the gaseous constituent to be removed by passing the gaseous stream
through the liquid.
44
5.1.2.1. Double alkali process
Absorption of CO2 by scrubbing with a caustic aqueous solution is a
familiar industrial process. The double alkali process involves scrubbing the gas
stream with an aqueous solution of sodium hydroxide in a packed column. The
CO2 is stripped from the gas by the following chemical reaction:
2NaOH + CO2 Æ Na2CO3 + H2O
The bulk of the treated gas is filtered by roughing and by high efficiency
particulate air filters before being released to the environment.
At appropriate intervals the spent scrubbing solution containing mainly
Na2CO3 is removed from the column to a mixing tank in which calcium
hydroxide is added to the solution, causing the carbonate to precipitate as
calcium carbonate:
Na2CO3 + Ca(OH)2 Æ 2NaOH + CaCO3
The solution and precipitate are then pumped to a filtration system,
where they are separated. The sodium hydroxide filtrate is cycled back to the
absorption column, while the calcium carbonate cake is removed for final
disposal as solid radioactive waste by incorporation into cement. The amount
of Ca(OH)2 used to precipitate the CaCO3 may have to be metered very
carefully to avoid any excess, which would cause saturation of the filtrate with
Ca(OH)2. The calcium hydroxide would in turn react in the packed column to
form CaCO3, which could plug the packing.
Removal of CO2 by caustic scrubbing involves absorption accompanied
by a chemical reaction. Experience indicates that the use of a 2M sodium
hydroxide solution and maintenance of a CO2 to carbonate conversion of
15_25% will generally yield an optimum absorption of CO2 [13].
5.1.3.
Physical absorption
5.1.3.1. Gas absorption by wet scrubbing
Carbon dioxide can be removed from a gas stream by contacting the
stream with water in a counter-current flow packed column. The process
depends on normal gas absorption principles.
By studying the solubility of CO2 in water in conjunction with the design
procedure for packed columns, it has been found that a large column is needed
to attain almost complete removal of the CO2. Space limitations in nuclear
45
power plants and reprocessing facilities make this method less feasible than
other methods.
5.1.3.2. Ethanolamine scrubbing
Gas scrubbing with ethanolamine (HOCH2CH2NH2) is one method for
the removal of CO2 from gaseous streams and involves absorbing the CO2 into
an ethanolamine solution at ambient temperatures in a scrubbing tower; the
solute is then steam stripped back out of the scrubbing solution in a second
contactor. Such a process removes CO2 from one gas stream and produces
another gas stream somewhat richer in CO2.
The particular operating problem of ethanolamine is its oxidation to
corrosive oxalic acid and glycine, which has been experienced in the gas
industry. This would possibly be amplified in a radiochemical application. A
product solidification technique would be required in addition to the
ethanolamine scrubbing.
5.1.3.3. Absorption in a fluorocarbon solvent
Carbon-14 as CO2 is quite soluble in liquid refrigerant-12 (R-12), dichlorodifluoromethane. Krypton, xenon and CO2 are considerably more soluble
than other volatile gases that may be in the gaseous effluent stream from a
nuclear power reactor, and are favourably temperature sensitive relative to the
others. Differences in relative solubility make gas separation possible. Since
lower temperatures increase the extraction efficiency, systems are designed to
operate at low temperatures (e.g. –17°C).
Two such processes have been developed for spent fuel reprocessing
plant off-gases: one at the ORNL and one at the Kernforschungszentrum
Karlsruhe, Germany [12]. Both processes are intended for the retention of 85Kr.
However, the solubility of CO2 in R-12 is greater than that of krypton, and the
possibility therefore arises of removing 14C from the off-gas stream in the same
plant as used for krypton removal.
The ORNL design originally consisted of an absorber, fractionator and
stripper, and was later modified into a single column operation containing the
three process operations. The product from the stripper column is a krypton–
xenon–CO2 mixture, and further separation is possible by cold traps. When
operating in the single column mode, concentration peaks for the various
components were observed at various locations along the column, and selective
removal of the enriched fluids was deemed possible.
The process developed at Karlsruhe used only two columns, each
consisting of absorbing, fractionating and stripping sections. Xenon and CO2
46
removal and enrichment are accomplished in the first column, and krypton
removal and enrichment in the second column [10]. However, CO2 removal
from a krypton–CO2 gas mixture may result in contamination of the CO2
fixation product by 85Kr and may therefore complicate waste disposal.
5.1.4.
Physical adsorption on an active surface
The capture of CO2 by fixed bed adsorption using molecular sieves would
require a pre-absorption step to dry the feed gas, removing essentially all the
water. A widely used molecular sieve that effectively removes CO2 is sodium
zeolite. The bed being loaded would have to be maintained between –75 and
_78°C to achieve good adsorption. Bed temperature is the most important
parameter for loading considerations.
The regeneration of a loaded bed is accomplished by heating it to a
temperature between 150 and 350°C, then passing a purge gas through the bed
at the elevated temperature to remove CO2. The purge gas would have to be
passed through the double alkali process to convert the 14C to a stable waste
form for storage and eventual disposal.
The type 13X molecular sieve is one of several types that can be used to
remove CO2. With proper bed regeneration, extremely high removal
efficiencies with decontamination factors greater than 100 are possible [10].
Mineo et al. [48] studied the performance of natural mordenite, hydrogenated mordenite and modified hydrogenated mordenite for the removal of
14
CO2 from dissolver off-gases. They concluded that the modification of hydrogenated mordenite by NaOH resulted in a larger adsorption capacity compared
with the other adsorbents. The capacity was found to decrease in the presence
of 1% NOx.
5.1.5.
Other methods
It has been suggested that the reaction of CO2 with magnesium at 600°C
could be used for 14C fixation [12]:
14
CO2 + 2Mg Æ 2MgO + 14C
The solid carbon product is expected to be stable and suitable for direct
immobilization and disposal.
Sakurai et al. [49] used a dry method to decompose 14CO2 into elemental
14
C through its reaction with H2 using microwave discharge. The reaction
produces CO as an intermediate, and proceeds in two steps:
47
CO2 + H2 Æ CO + H2O
CO + H2 + Cn Æ Cn+1 + H2O
where Cn denotes the carbon already deposited on the wall of the discharge
tube. The results showed that microwave discharge for a CO2–H2 mixture in
reduced pressures (~0.67 kPa) reduced CO2 into carbon. The reduction of CO2
into CO proceeded fast, and there was a low rate of conversion of CO into
carbon.
The recovery of 14C from an irradiated graphite moderator of a decommissioned Calder Hall type power reactor has been studied in Japan [50]. The
recovery process consists of three main steps: graphite combustion, 14C
enrichment and 14CO2 immobilization (Fig. 7). Carbon isotope exchange
between CO2 and carbamate was employed for the 14C enrichment. The
process can be described by the following isotope exchange reactions:
14
14
12
14
–1
CO2 + R2N12CO–1
2 ´ CO2 + R2N CO2
12
12
14
–1
CO2 + R2NH CO2 ´ CO2 + R2NH CO2–1
The performance of 14C separation was calculated under the conditions
that the concentration of 14CO2 in the stripped flow is less than the environ-
Irradiated
graphite
Calder Hall type
power reactor
CO2
Combustion
process
Carbon isotope
separation process
Depleted
14
CO2
less than
environmental
standard
CO2
Enriched
14
CO2
Disposal after
cementation
Precipitation
by alkaline
Released to
atmosphere
FIG. 7. Carbon-14 recovery process from irradiated graphite.
48
mental standard (3 Bq/cm3) and that the stripped flow corresponding to 99% of
the feed CO2 is released into the atmosphere. The process dimensions, which
were calculated for the recovery of 14C from 1600 t of graphite used in the
Japanese Calder Hall type reactor Tokai 1 (166 MW(e)), are estimated to be
5.2 m in diameter and 16 m in height for continuous operation when using the
conventional dibutylamine (DBA) octane solution system. If diethylamine
(DEA) octane is used as a working fluid and the process is operated at higher
pressures (0.2 MPa), the column dimensions can be decreased to 3.2 m in
diameter and 5.7 m in height. If, however, the process is operated at 0.3 MPa
with 4M of DEA, then the exchange column can be designed compactly with
dimensions of 3.3 m in diameter and 6.7 m in height. For this compact design,
the process can be operated batchwise at the rate of four batches per month.
These results suggest that the CO2 carbamater method is applicable for the
recovery of 14C from irradiated graphite. In particular, the batch operation, in
which the exchange column can be operated simply in a total reflux mode, is an
attractive option.
A sophisticated off-gas treatment system has been proposed to enrich
14
CO2 from graphite incinerator off-gas (Fig. 8). This system, which is based on
the separation nozzle process (i.e. the Helicon process), has been used in the
past for uranium enrichment and the separation of gaseous isotopes. The
treatment system consists of a primary separator for the separation of light
gases such as nitrogen, oxygen and steam from the relatively heavy gases 12CO2
and 14CO2, and a secondary separator to separate 14CO2 from 12CO2, based on
their density difference (i.e. the density of 14CO2 is 2.05 g/L and of 12CO2 is 1.96
g/L). The enriched stream is expected to contain approximately 80% of the
14
CO2 of the feed stream. This enriched stream will be neutralized by barium
hydroxide to form insoluble barium carbonate [11].
In general, 14C separation methods are costly and require a high energy
consumption. Applications of these separation technologies may therefore be
limited by their high cost.
5.2. REMOVAL OF 14C FROM LIQUID WASTE
Carbon-14 in aqueous liquid waste is usually in chemical forms of
carbonate and/or bicarbonate, depending on the solution pH. Conventional
water treatment processes, such as ion exchange, can therefore sufficiently
remove 14C as HCO3–1 /CO3–2 from liquid streams.
The amount of 14C in liquid waste produced during reactor operations is
relatively small (Table 18). The active liquid waste treatment system installed in
a reactor is designed for the removal of other radiochemicals (e.g. 60Co and
49
137
Cs) and some toxic chemicals (e.g. boron, PO4–3 and organics). Carbon-14 in
liquid water is expected to be removed by this treatment system. In general,
when reporting the discharge water quality from a nuclear facility, 14C is
grouped with other beta emitters as gross beta activity in which tritium and 90Sr
may be the dominant species.
At fuel reprocessing plants the principal liquid stream containing 14C is
the spent caustic solution from the off-gas scrubbing. The 14C (e.g. as Na2CO3)
can be removed by contacting with compounds of calcium or barium to form
more stable carbonate compounds. The solid carbonate is separated from the
liquid and is usually solidified by a binding material (e.g. cement) for long term
storage and/or disposal.
Inlet flux
Seventeen depleted stages
Twenty-four enrichment stages
Enriched
flux 14CO2
Depleted
flux 12CO2
Enriched 14CO2 2.4 m3/h
Ba(OH)2
BaCO3
Twenty-four
enrichment stages
Stack
CO2 224 m3/h
Seventeen
depleted stages
Primary cyclon
Depleted
12
CO2 221 m3/h
Light gas exhaust
FIG. 8. Processing off-gas from irradiated graphite incineration for 14C recovery.
50
5.3. SEPARATION OF TRITIUM FROM SPENT FUEL
Tritium removal from gas streams and liquid streams can be uneconomic,
especially when the tritium concentration is low. If the goal is the removal or
recovery of tritium, it is desirable to apply the removal technologies before
additional dilution takes place. However, the common practice in many
instances, particularly in reprocessing plants and LWRs, is to allow dilution to
take place within the plant processes, followed by further dilution upon release
to the environment.
Tritium can be separated from spent fuel prior to the dissolution step in
reprocessing plants. The released tritium can then be removed from the
evolved gas by means of tritium removal methods. Two technologies have been
developed for the separation of tritium as a gas: voloxidation and pyrochemical
processing.
5.3.1.
Voloxidation
This process was developed as a means of separating tritium from
irradiated reactor fuel prior to dissolving the fuel in nitric acid for reprocessing
operations. The voloxidation process depends on the oxidation of UO2 to U3O8
in order to break down the crystal lattice and release gases from it. Tritium
release efficiencies of greater than 90% are expected from the fuel with this
method, which exposes the fuel to 450–500°C for several hours in a rotary kiln.
The evolved gas is passed through a catalytic converter to ensure that all the
tritium has been converted to HTO, and then it is removed as water in solid
absorbents. A large portion of the tritium left in LWR fuel elements is bonded
in the zircaloy cladding and will remain with the cladding during fuel dissolution. This process may not work for ThO2 fuels, since there is no higher oxide
of thorium. Metal fuels may also be difficult to treat in this manner because
heat generation rates may be difficult to control during oxidation [51].
5.3.2.
Pyrochemical processing
Pyrochemical processing involves the application of high temperatures.
This approach is considered to be a head end treatment in the reprocessing
plant and involves two steps:
(a)
(b)
Chopping and/or cladding removal;
Treatment of the oxide fuel to effect the release of volatile fission
products, including tritium.
51
Decladding by melting has been studied for stainless steel clad fuels; in
principle, it could be used for zirconium clad fuels. Unlike the voloxidation
process, in which UO2 is oxidized to U3O8, in this process the second step will
normally be a reduction step in which uranium and plutonium metals are
produced by reduction in the presence of a molten salt. A mixture of zinc,
calcium and magnesium or another reductant alloy and a salt such as calcium
chloride is heated with the oxide fuel at about 800–900°C. The principal
reaction is:
UO2 + 2Ca Æ 2CaO + U
The salt is separated from the metal phase. After distilling the magnesium
and zinc for recycling, the uranium and plutonium are sent to the acid dissolver
for solvent extraction processing of the aqueous solution [52].
Fission product tritium and noble gases are both released during the
decladding and reduction steps and are handled as in the voloxidation process.
5.4. REMOVAL OF TRITIUM FROM GAS STREAMS
In the off-gases from nuclear facilities, tritium can be present as HT or
HTO. Removal methods for HT and HTO are briefly discussed below.
5.4.1.
Removal of HT
There are non-oxidative processes for the removal of HT that employ
hydrogen getters. These getters typically consist of zirconium alloys and
compounds. However, the selection of the appropriate getter depends greatly
on the other constituents of the gas. The fact that most getters will not tolerate
oxygen in the feed stream greatly limits the usefulness of getters for most
effluent streams.
5.4.2.
Removal of HTO
The typical processes for removing tritium from gas streams involves
conversion of HT to HTO followed by the removal of the HTO. A number of
methods exist for removing HTO vapour from gaseous streams, depending on
the degree of removal required.
52
5.4.2.1. Molecular sieves
A very high degree of HTO removal can be achieved through the use of
molecular sieves. Lakner et al. [53] reported that up to 47 cm3 of HTO per gram
of molecular sieve (at standard temperature and pressure) could be obtained.
Absorption on molecular sieves is a reversible process and allows for the regeneration of the molecular sieve and the recovery of the tritium as water vapour
by heating to moderate temperatures (e.g. 250°C). In practice, recovery of this
tritium is not usually determined to be cost effective. The molecular sieves
containing the HTO are typically stored, disposed of as low level waste or the
water is transferred to another medium for disposal.
5.4.2.2. Dehumidification
Packed bed or plate columns can be used to counter-currently contact the
gas stream with a chilled water stream. The degree of water removal from the
system is not great, but the performance may be improved by an exchange of
water vapour between the gas and liquid phases. This technique would only be
used if the intention is to generate a dilute tritium stream for direct discharge.
5.5. REMOVAL OF TRITIUM FROM LIQUID WASTE
In general, the volumes of liquid waste generated from the fuel cycle can
be reduced by recycling; this is true for both reactors and reprocessing plants.
This, however, raises the concentration of the tritium in the waste. Methods for
removing tritium from liquid waste provide an alternative to the control of
tritium emissions and personnel exposure.
The most commonly practised application of tritium removal from liquids
is the detritiation of heavy water. In this application, the detritiation and
recycling of the heavy water is much more economic than the production of
new heavy water.
A number of techniques have been developed, primarily for tritium
removal or the detritiation of heavy water (for HWRs). These techniques
include electrolysis and catalytic exchange and fractional distillation. Tritium
removal from light water, however, requires larger stripping and recovery
factors than for heavy water upgrade. To be capable of obtaining the required
high recovery, a feasible tritium removal process will require a high isotopic
separation factor.
The hydrogen–water chemical exchange process has the required high
isotopic separation factor and is suited to a water feed. In addition, the
53
technical feasibility of hydrogen isotopic enrichment through the exchange
between hydrogen and water is established.
Previous IAEA reports [23, 54] have described the state of tritium
technologies at the time of their publication (1981 and 1984, respectively),
whether speculative, in development or in use. Table 28 provides a summary of
these technologies and their current status.
5.5.1.
Tritium enrichment
The process of detritiating water consists of three main steps [55]:
(a)
A front end step that exchanges the tritium to the less toxic hydrogen
phase. This can be performed either through chemical exchange in the
presence of a platinum based catalyst:
HTO + H2 ´ H2O + HT
or through the decomposition of water:
2HTO Æ 2HT + O2
(b)
(c)
A back end process that concentrates the tritium in the hydrogen phase.
Cryogenic hydrogen distillation is usually preferred for this step, since
other potential processes, such as gas chromatographic methods, are not
economic for large scale applications.
A means of stabilizing the concentrated tritium and storing it safety.
Uranium metal is used if storage is temporary; titanium is mainly
employed for long term storage.
Several technologies exist for the front end step. The most industrially
proven process is vapour phase catalytic exchange (VPCE), which was
developed and patented by the Commissariat à l’énergie atomique (CEA) in
France. The VPCE process normally comprises three to eight stages, each stage
consisting of a water evaporator, superheater, catalyst bed and condenser–
separator. Since a conventional platinized catalyst is used, superheating is
required to minimize steam condensation within the catalyst pores, which
would inhibit isotope exchange. Water is fed to the first stage, where it is
evaporated and mixed with cold deuterium gas returning from the condenser–
separator of the subsequent stage. This gas–vapour mixture is superheated to
200°C prior to entering the catalyst bed, where tritium is exchanged and equilibrated. The deuterium gas, somewhat enriched in tritium, is separated from
54
TABLE 28. STATUS OF TECHNOLOGIES AND PROGRAMMES DESCRIBED IN PREVIOUS IAEA PUBLICATIONS ON TRITIUM AND TRITIUM WASTE
Process
Technology
Publication
Status in publication
Current status
Heavy water
detritiation
Water distillation
TRSa 203 [23]
Has been used to remove tritium
from D2O in HWRs at low
throughput
Used in ‘upgraders’ at HWRs to
remove light water from heavy
water
TRS 234 [54]
Offered as technology that does not
require handling of hydrogen gas,
but is energy intensive
TRS 203 [23]
Offered as tritium enrichment
technology to be used in
conjunction with VPCE, LPCE,
combined electrolysis and catalyst
exchange (CECE) or direct
electrolysis
Used in conjunction with LPCE to
detritiate heavy water from all
operating CANDU reactors in
Canada
TRS 234 [54]
Has operated with a VPCE process
in Grenoble, France, since 1972
Continues to operate in Grenoble,
France
Ontario Hydro built a compact
cryogenic distillation system,
which operated at the Princeton
Plasma Physics Laboratory in the
mid-1990s
TRS 203 [23]
Successfully demonstrated with
cryogenic distillation in Grenoble,
France
Continues to operate in Grenoble,
France
Cryogenic
distillation
Vapour phase
catalytic exchange
(VPCE)
55
56
TABLE 28. STATUS OF TECHNOLOGIES AND PROGRAMMES DESCRIBED IN PREVIOUS IAEA PUBLICATIONS ON TRITIUM AND TRITIUM WASTE (cont.)
Process
Technology
Publication
Status in publication
Current status
TRS 234 [54]
Projected to be used by Ontario
Hydro in tritium removal systems
Ontario Hydro selected LPCE
instead of VPCE for tritium
removal
TRS 203 [23]
Proposed as the preferred method
for a Canadian detritation process;
only demonstrated at the laboratory
scale
AECL modified the process at
Chalk River Laboratories to use
CECE instead of LPCE; it was
started up as a demonstration
process in 1999
TRS 234 [54]
Projected to be used by AECL at
Chalk River Laboratories
Combined
electrolysis and
catalytic exchange
TRS 203 [23]
Exchange process developed by
AECL; one of the candidates for
future detritiation systems
Volatilization
TRS 203 [23]
Speculative process, not yet
developed; drawbacks include high
cost and difficulty controlling the
exothermic reaction
Liquid phase
catalytic exchange
(LPCE)
Tritium
recovery as
gas at fuel
reprocessing
plants
CECE system was built at the
Mound Facility, Ohio, USA, to
detritiate light water from
glovebox cleanup operations; it
operated in the 1980s
TABLE 28. STATUS OF TECHNOLOGIES AND PROGRAMMES DESCRIBED IN PREVIOUS IAEA PUBLICATIONS ON TRITIUM AND TRITIUM WASTE (cont.)
Process
Tritium
recovery from
water at fuel
reprocessing
plants
a
Technology
Water distillation
TRS: Technical Reports Series.
Publication
Status in publication
TRS 234 [54]
Technology not yet fully developed;
additional drawback is the
behaviour of remaining volatiles
(carbon, ruthenium, krypton,
iodine)
TRS 203 [23]
Has been applied at an LWR fuel
reprocessing plant
TRS 234 [54]
At present there has been no
attempt to trap or retain tritium at
reprocessing plants
Current status
57
water in the condenser and fed to the back end cryogenic distillation system.
The tritium depleted water flows to the evaporator of the subsequent stage.
With the development of a wetproofed catalyst by AECL [56], the
catalyst isotopic exchange process can be greatly simplified to a packed catalyst
column, with liquid entering from the top and trickling down the column in
counter-current contact with deuterium gas that is flowing up. Wetproofing
allows the catalyst to operate in liquid or very humid environments. This
process, which operates at 50°C, is referred to as liquid phase catalytic
exchange (LPCE), and has been patented by AECL. India, the Republic of
Korea, the Russian Federation and the USA also report that they have
developed a wetproofed catalyst.
A third option for the front end step is combined electrolysis and catalyst
exchange (CECE). This process has also been developed and patented by
AECL. Unlike the other options, this process performs some pre-enrichment
of the tritium stream before sending the deuterium to the cryogenic distillation
system. The process consists of LPCE columns and electrolytic cells. The
catalyst column is divided into stripping and enriching sections. HTO is fed to
the top of the enriching section, in which it picks up tritium from the electrolytic deuterium produced from the water leaving the bottom of the column. A
side stream of tritium enriched deuterium gas is taken from the electrolysis cell
as feed for the cryogenic distillation system and is returned detritiated higher
up in the exchange column. The deuterium gas flowing through the stripping
section is stripped of tritium by the detritiated reflux water formed by catalytic
recombination with oxygen produced in the electrolytic cells. The detritiated
product water is also drawn from the recombiner.
In the late 1980s the USDOE’s Mound Facility in Miamisburg, Ohio,
USA, also operated a CECE process for light water detritiation.
The last technology option frequently considered for the front end step is
electrolysis, in which water is completely electrolysed to hydrogen and oxygen.
The hydrogen gas is fed directly to the cryogenic distillation system, where it is
detritiated and then sent to a catalytic or frame type recombiner to re-form
with the oxygen from the electrolytic cells, to produce the detritiated water
product. This option is power intensive and expensive.
5.5.2.
Other tritium removal technologies
Other options for tritium phase conversion include water distillation,
laser isotope separation and chemical decomposition, but for various reasons
none have been considered to be serious contenders [55].
Molecular Separations, Inc., (MSI) [57] has developed a new approach to
the separation of diluted HTO from light water. The process uses modified ion
58
exchange resins to selectively adsorb the heavier water molecules, in particular
HTO, relative to light water. MSI conducted pilot scale tests using feeds from
an operating nuclear power plant and groundwater from USDOE sites. The
reported cost of this process is approximately $2.5/L, with an average of a 70%
reduction in tritium content.
Owing to handling and safety considerations, tritiated organic liquid
waste is usually converted into HTO and CO2 prior to further treatment.
Catalytic and thermal oxidation are two well established technologies. Other
options are wet oxidation based on chemical, electrochemical, biochemical or
photochemical principles [58].
Ontario Hydro Technologies has developed a compact, low inventory
cryogenic distillation system for use at the Princeton Plasma Physics
Laboratory in Princeton, New Jersey, USA. While the system was used for the
cleanup of the exhaust from the Tokamak Fusion Test Reactor, the same
cryogenic distillation technology is also suitable for the isotope separation step
of a tritium recovery process [59].
In 1998 AECL completed a study, commissioned by the USDOE, which
examined the detritiation of light water containing very low concentrations of
tritium. Of the nine technologies examined, six were judged to be uncompetitive or impractical: water distillation, electrolysis, bithermal ammonia
hydrogen, monothermal ammonia hydrogen, laser isotope separation and
water permeation. The three leading processes all include chemical exchange:
CECE, Girdler sulphide and bithermal hydrogen–water exchange [60].
6. ANALYTICAL AND MONITORING METHODS
Monitoring of 14C and tritium is of key importance to the managers and
operators of nuclear facilities. Accurate and reliable methods to measure the
concentrations of the contaminants to be released are necessary for proper
emission system operation and, more importantly, to demonstrate to the public
and regulators that emissions and nuclide concentrations in stored waste are
within acceptable limits.
Detection of surface contamination is also of concern for safe plant
operation; however, this type of monitoring is more within the scope of
radiation protection than waste management and is therefore not covered in
this report.
Effluent monitoring systems are intended to measure radionuclides in
airborne and liquid effluents before discharge to the environment, while
59
environmental monitoring systems are intended to measure levels of radionuclides in selected environmental media: the two systems are complementary.
Effluent monitoring is always required if radiologically significant amounts of
radioactive contaminants are being released from a facility or if there is a
potential for radiologically significant unplanned releases. Environmental
monitoring may also be required if the potential releases of radioactivity could
result in a significant dose to critical groups or to the whole population.
Environmental monitoring provides a more direct assessment of the levels of
radionuclides to which members of the public are exposed. Information on the
objectives and design of environmental monitoring programmes and effluent
monitoring programmes are given in Refs [61, 62].
6.1. CARBON-14 MONITORING SYSTEMS
6.1.1.
Carbon-14 sample collection
6.1.1.1. Air samples
Techniques for sampling 14CO2 in air can be either active or passive.
Active techniques for air sampling are adapted from the methods developed
for sampling and monitoring 14C in reactor stack effluents. In most of these
methods air is pulled through a bubbler containing sodium hydroxide or a tube
filled with a molecular sieve. The flow rate must be kept low to allow sufficient
time for the CO2 to be trapped. Concentrations of NaOH used for sampling are
in the range 0.8–4M.
14
CO, 14CH4 and other hydrocarbons present in air samples can be
measured separately by oxidizing the gas effluent from the first bubbler (to
trap 14CO2) and capturing the 14CO2 produced in a second NaOH bubbler [63–
65]. Alternatively, air samples can be oxidized first, in which case they become
an integral part of the total 14CO2 measured [38, 66]. The different carbon
compounds are oxidized under different conditions. CuO [38, 67] and Pd/Pt
[63, 68, 69] are the catalysts most commonly used for oxidation processes, at
temperatures ranging from 400 to 800°C.
Passive samplers can be liquid or solid. Liquid passive samplers are used
at AECL’s Chalk River Laboratories. Four-molar NaOH solution is put into
plastic trays, covered with nylon mesh and set on stands attached to pointed
metal rods, which can be driven into the ground; they are protected from rain
and left unattended for one to two weeks [70]. An alternative sampler with
filter paper wetted with 0.8M of NaOH solution has been used at AECL’s
Chalk River Laboratories to absorb 14CO2. This system is only good for short
60
term measurements because the NaOH solution would saturate if longer term
sampling were attempted. Other passive samplers, such as those used by Otlet
et al. [71], use a variant of this passive sampling method.
Solid adsorbents have obvious advantages for handling and transport,
and have been used for monitoring reactor stack gases [38, 63]. The use of
NaOH pellets as the absorbent in a passive sampler has been examined at
AECL’s Chalk River Laboratories. A single pellet (100–200 mg) provides more
than enough capacity for sampling periods of up to a week.
6.1.1.2. Water samples
Carbon dioxide can be recovered directly from water samples by acidification and purging with an inert gas (e.g. nitrogen or helium) to transfer the gas
into a collector. In general, bicarbonate–carbonate in the samples is recovered
as barium carbonate on the addition of barium chloride and NaOH to raise the
pH to pH10 [70].
Methane in water samples can be recovered by the purging technique.
The CO2 evolved from the samples is captured in liquid nitrogen traps, while
the CH4 proceeds to a furnace, in which it is oxidized to CO2. This CO2 is
captured in two more traps for later analysis [70].
6.1.1.3. Vegetation and soils
Samples of leaves and needles are collected in new polyethylene bags;
these samples are rinsed in distilled water and dried overnight at approximately
80°C prior to storage in glass sealers [70].
Soil samples can be obtained by driving 10 cm lengths of aluminium
tubing into the ground. As soon as the core tube is withdrawn, both ends
should be capped, to maintain the integrity of the sample until it can be
extruded and sliced into sections. A subsample of each section is taken to
determine the percentage of water and organic carbon. The remainder of the
sample should be stored frozen for later analysis, to prevent bacterial or fungal
activity [70].
6.1.2.
Carbon-14 sample preparation
Since 14C occurs in both liquid and gaseous streams in reactors and
reprocessing plants, analytical methods are required for a variety of sample
types. The low energy of the beta radiation and the absence of gamma
emissions in 14C decay, plus the fact that other radioactive gases are always
61
present, make it essential to separate the 14C species before any measurement
is made [72].
The separation procedure will vary depending on the 14C levels and the
impurities expected; for example, Kunz [73] described procedures for
separating gases for internal gas proportional counting from various nuclear
facilities. The gases separated included argon, krypton, xenon, H2, CH4, CO2
and water vapour.
In reprocessing plants it will be necessary to use more elaborate
separation methods than those used for the samples taken from a reactor
environment. Braun et al. [32] described a scheme for the separation of 14C as
14
CO2 from dissolver off-gases. After uranium fuel dissolution, the off-gas
comprises air as carrier gas, aerosols, iodine, NOx, krypton, xenon, CO2, water
and hydrocarbons. The NOx present in the off-gas sample is extracted by
scouring with an aqueous solution to produce HNO3; the aerosols and iodine
are removed by various filtration processes, and any hydrocarbons present are
converted to CO2 by oxidation.
In general, the sample preparation method will depend on the nature and
physical state of the sample, and also on the measurement technique adopted.
In all cases, however, the 14C in the sample is at some stage converted into CO2.
For air and water samples, CO2 is released from either NaOH solution or
precipitated BaCO3 on neutralization of the sample with acid. Carbon dioxide
adsorbed on a molecular sieve is usually removed by heating and eluting with
nitrogen or helium gas. Solid samples are usually combusted to CO2 if they are
combustible [70].
Following its release, CO2 prior to 14C catalysis can be collected as a gas,
either in a gas bag or a cold trap, and subsequently:
(a)
(b)
(c)
(d)
(e)
(f)
(g)
(h)
62
Purified by capturing on activated charcoal [74];
Absorbed on solid Ca(OH)2 [38];
Converted to CH4 by catalytic hydrogenation [75];
Converted to C2H2 by synthesis of LiC2 and its subsequent hydrolysis
[70];
Converted to benzene by trimerization of C2H2 [70, 76];
Absorbed directly into an amine such as Carbo-Sorb [70];
Reduced to graphite with magnesium metal or H2, using iron or cobalt as
a catalyst [77];
Trapped in NaOH [78], followed by precipitation as BaCO3 or CaCO3 or
by direct release into a scintillation cocktail for measurement.
6.1.3.
Analytical methods for 14C
There have been many reviews of 14C analytical techniques (e.g. Refs [21,
67, 69, 79]). Brief summaries of the available analytical techniques are given in
Sections 6.1.3.1–6.1.3.3.
6.1.3.1. Gas counting
Gas proportional counters combine the advantages of high counting
efficiency with pulse height discrimination. Proportional tube volumes range
from 5 mL to 7.5 L. Methane as the counter filling has been shown to provide
better resolution and gas gain stability, whereas CO2 is easier to prepare and
handle but must be purified from the sample [75]. Efficiencies greater than
90% and backgrounds of 0.1 cpm are at present achievable for 14CO2 [70].
6.1.3.2. Liquid scintillation counting
As a result of large improvements in performance, liquid scintillation
spectrometry is now the measurement technique employed for 14C at most
laboratories. Efficiencies of 80% and backgrounds of approximately 5 cpm are
possible on a routine basis; backgrounds of less than 0.5 cpm are achievable
with anti-coincidence shields. The nature of the scintillation cocktail chosen
will depend on the final sample form, vials to be used, etc. Many reviews of this
technique can be found [80–82].
6.1.3.3. Accelerator mass spectrometry
Radiocarbon measurement using single atom counting by accelerator
mass spectrometry provides high precision measurements with extremely small
samples [83]. At AECL’s Chalk River Laboratories samples were prepared by
reduction of CO2 to graphite by reacting the gas with red hot magnesium
turnings, using iron powder to facilitate packing the purified carbon into the
target cones. Current reduction techniques use H2 with an iron or cobalt
catalyst [77]. This method has been used mostly in the dating of very small
amounts of material, but the cost per measurement at dedicated accelerator
mass spectrometry laboratories is competitive with that of other methods [70].
6.1.4.
Carbon-14 monitoring methods
Monitoring systems for 14C used in nuclear facilities have been developed
based on the sampling methods discussed above; for example, Kunz [84]
63
developed a continuous sampler to monitor the 14C levels released from an
LWR. The system consisted of an oxidizing furnace, solid absorbents for H2O
and CO2, a metering valve, a flow meter and a pump. The sampled gas was
passed at 600°C through a tube furnace containing pellets of palladium on
alumina and platinum on alumina to oxidize all carbon species (CH4, C2H6, CO,
etc.) to CO2. From the furnace, the gas flowed through an absorbent to remove
water vapour and then through a cartridge containing an absorbent to capture
the CO2. Commercially available solid absorbents were used for both the water
vapour (Drierite) and CO2 (Ascarite). The absorbent cartridges were changed
at regular intervals.
To measure the 14C in the sample, CO2 was evolved from the Ascarite by
acidification. A portion of the recovered CO2 was chromatographically
purified, and the 14C concentration was measured in a gas proportional
counter.
Kabat [38] developed a similar system, in which the gaseous organic
compounds and CO are first oxidized to CO2 by catalyst (CuO) combustion;
CO2 is then absorbed on solid Ca(OH)2 at elevated temperatures. Subsequently, 14CO2 is liberated by thermal decomposition of Ca14CO3 and is
scrubbed in an NaOH solution and measured by liquid scintillation counting.
6.2. TRITIUM MONITORING SYSTEMS
6.2.1.
Air samples
Tritium may be present in air as tritium gas or as tritium oxide. Tritium
gas cannot be trapped as easily as tritium oxide, and it must first be converted
into tritium oxide by passing the air stream over heated copper oxide wool [62].
Sampling of air by a tritium sampler and subsequent measurement of the
accumulated tritium activity by liquid scintillation counting is a simple, reliable
and inexpensive method of monitoring for HTO vapour in air [85]. In most
sampling methods air is passed through a bubbler containing water or a tube
filled with a molecular sieve or silica gel. Studies (e.g. Ref. [86]) have shown
that the collection efficiency of a single water bubbler exceeds 95% for short
term sampling (~10 min), even at an air flow rate as high as 10 L/min. An
ethylene glycol and water mixture is used in place of water in the bubbler to
minimize evaporative losses in longer term sampling or for use under freezing
conditions.
Many different types of gas sampler have been developed and used for
measuring very small quantities of tritium in very large volumes of gas. These
samplers are used to measure quantities of tritium released through a facility’s
64
stack and for environmental monitoring at a site. Stack exhaust gas monitoring
systems generally use an ionization chamber to measure tritium in stack gases
and a gas sampler to measure the extremely low levels of tritium that cannot be
measured by an ionization chamber.
Most of the stack gas samplers are of a similar design as the ethylene
glycol sampling system developed at Mound Laboratories, Ohio, USA. A
commercial version of this system is now available. In this system, a sample of
gas from the stack is circulated through six ethylene glycol bubblers in series.
The first three bubblers remove tritium in the form of HTO, DTO and T2O.
The gas stream is passed through a heated catalytic reactor, in which tritium in
the form of HT, DT, T2 and CHxTy is cracked and oxidized to form water. This
sample is then passed through three more ethylene glycol bubblers to remove
the tritium gas, which is now in the form of water. After a period of a few hours
to days, a sample of the ethylene glycol from each bubbler is removed and
counted using a scintillation counter to determine the quantity of tritium in
each bubbler. The tritium recovered from the first three bubblers is proportional to the tritium in liquid form contained in the stack gases, and the tritium
recovered in the last three bubblers is proportional to the quantity of tritium in
gaseous form contained in the exhaust gases.
Owing to the extremely small quantity of tritium contained in the
atmospheric gases surrounding a tritium facility, environmental gas samplers
use higher flow rate sampler systems than those required for stack monitoring,
and, in general, collect the water on molecular sieve traps. The water collected
on the molecular sieve traps is then recovered from the trap and the tritium
concentration of the gas passing through the trap is calculated from the tritium
concentration of the collected water, the gas flow rate through the trap and the
sampling time.
The use of solid adsorbents such as a molecular sieve or silica gel is particularly suited to long term sampling to establish long term tritium contamination levels in the workplace. The tritium oxide accumulated in the sampler is
extracted by heating and the water collected is analysed for tritium by liquid
scintillation counting.
Both the liquid and solid samplers discussed above require the use of an
air pump as well as a means of measuring and controlling the sampling flow at
a constant rate over the sampling period. A passive diffusion sampler has been
developed in Canada [87, 88]. The sampler consists of a standard 20 mL scintillation vial with a vial lid modified to contain a stainless steel insert with an
accurate diffusion orifice in the centre. HTO vapour diffuses through the
orifice and is trapped by an appropriate HTO sink, such as de-ionized water.
The rate at which HTO in air enters the vial is determined by the diffusion rate
of the diffusion orifice. At the end of the sampling period, 1 mL of water is
65
added to the vial; after an equilibration period of about 1 min, 15 mL of liquid
scintillation cocktail is added and mixed with the contents of the vial, which is
then measured for tritium in a liquid scintillation counter. The diffusion
sampler is less expensive and easier to service than the active tritium samplers
discussed above. In addition, the diffusion sampler does not require an air
pump or power supply. These features make diffusion samplers very attractive
for routine outdoor tritium in-air sampling. Field evaluations of this passive
tritium sampler have been performed at AECL’s Chalk River Laboratories and
at Ontario Power Generation’s nuclear power plants in Canada [89–91]. The
results have indicated that the passive sampler can be used to measure HTO
vapour in air at concentrations as low as 1 Bq/m3 over a month.
Since HTO is more toxic than HT (>10 000 times greater), it may be
desirable to know the relative amounts of each species. In the case of stack
monitoring, discrete samples of the stack effluent should be taken using active
samplers such as water bubblers. A technique for differential monitoring uses a
desiccant cartridge in the sampling line of an ionization chamber monitor. The
result is a measurement of the HT concentration. The total tritium concentration is determined without the cartridge. Subtraction of HT from the total
produces the HTO concentration [21].
Another technique uses a semi-permeable membrane tube bundle in the
sampling line to separate HTO from HT, which is routed to an HTO monitor.
After removing the remaining HTO with another membrane, the sampled air is
directed to an HT monitor. This technique is slower than the one using a
desiccant cartridge, but it can be adapted to measure tritium in the presence of
noble gases or other radioactive gases by adding a catalyst after the HTO
dryers, followed by additional membrane dryers for the HTO [21].
6.2.2.
Samples of liquids
Tritium in liquid samples (e.g. wastewater and groundwater) is almost
universally measured by liquid scintillation counting; however, the liquid must
be compatible with the cocktail. Certain chemicals in a liquid sample can
degrade the cocktail, others may retain much of the tritium and still others
result in a high degree of quenching, therefore the sample should be pretreated
prior to the counting. In addition, samples that contain peroxide or that are
alkaline may result in chemiluminescence, which can interfere with
measurement. Such samples should be neutralized before counting [21].
66
6.2.3.
Tritium analytical methods and monitoring systems
Several different types of instrument can be used to detect and measure
tritium in the operation of a facility [9, 62]. Examples and a discussion of such
instruments are given in Sections 6.2.3.1–6.2.3.10.
6.2.3.1. Ionization chambers
Fixed ionization chambers are the most widely used instruments for
measuring gaseous forms of tritium in laboratories and in process monitoring
applications. Such simple devices require only an electrically polarized
ionization chamber, suitable electronics and a method for moving the gas
sample through the chamber, such as a pump.
Tritium decays to 3He by the ejection of a beta particle. The beta particle
generated by the decay of tritium ionizes the surrounding gas. The number of
ions produced due to the loss of energy of the beta particle is a function of the
type of gas. A sample of gas is collected in the ionization chamber and the
ionization current is measured. The resulting chamber ionization current is
proportional to the quantity of tritium in the gas. The larger the measuring
chamber volume the higher the output current and the easier it is to measure.
However, as the volume of the chamber increases, the longer it will take to
obtain an accurate measurement. Chamber volumes typically range from a
tenth to a few tens of litres, depending on the required sensitivity [21]. Modern
electronic systems have solved most of the problems associated with measuring
small ionization currents in small volumes, and, as a result, the volume of
ionization chambers has been reduced over the years from 50 L down to 1 or 2
L. Most tritium measuring instruments use an ionization chamber.
Although most ionization chambers are the flow-through type that
requires a pump to move the gas, a number of facilities use open window or
perforated wall chambers. These chambers, which employ a dust cover to
protect the chamber from particulates, allow the air or gas to penetrate through
the wall to the inside chamber without the need for a pump. These instruments
are used as single point monitors to monitor rooms, hoods, gloveboxes and
ducts.
6.2.3.2. Proportional counters
Gas proportional counters can be used to measure the amount of tritium
contained in a gas. A sample of the gas to be monitored is mixed with a
counting gas and passed through a proportional counter tube, in which the
pulses caused by the decay of tritium are counted. Proportional counter
67
monitors can be used for most gas monitoring applications and are also used to
measure surface contamination.
6.2.3.3. Scintillation crystal detectors
Scintillation detector systems are used to measure the total mole per cent
of tritium in a sample of gas, independent of the chemical composition of the
tritium in the gas (HT, DT, T2 and CHxTy). A sample of the gas is introduced
into a measurement chamber at low pressure, generally less than a few torr
(approximately 1–5 kPa). The chamber contains a scintillation crystal, which is
exposed to the tritium as it decays. The light pulse produced in the scintillation
crystal is either counted or is used to produce a current that is proportional to
the mole per cent tritium contained in the gas sample. Crystal scintillation
detection is generally used to measure the mole per cent of tritium in gases
containing high concentrations of tritium.
6.2.3.4. Mass spectrometers
Magnetic sector, quadrupole and drift tube mass spectrometers are used
as analytical tools to measure the individual components that make up the gas
being measured. Mass spectrometers are generally used for assay and accountability or for scientific purposes. A sample of the gas to be measured is
introduced at low pressure (approximately 1–5 kPa) into a chamber and
ionized. The ions produced are then measured by a means that discriminates on
mass. The number of ions produced at each mass is measured and is proportional to the partial pressure of the component in the gas sample. Light isotope,
drift tube mass spectrometers require a large capital investment and a skilled
staff for operation, and, in some cases, may not be cost effective. Quadrupole
mass spectrometers and crystal scintillation detectors are much less expensive,
but still require operation by knowledgeable, well trained personnel.
6.2.3.5. Liquid scintillation counters
Owing to the need to measure removable tritium on surfaces and in the
body water of workers, almost all tritium facilities are equipped with or have
access to a liquid scintillation counter. If a scintillation counter is not available
on the site, the service can generally be purchased from a local firm. Liquid
scintillation counters are used to measure the quantity of tritium on surfaces, in
liquids and in dissolved samples. For removable surface contamination
measurements, a wipe of the surface to be measured is taken using dry filter
paper or a cotton swab. The filter paper or cotton swab is then placed in a
68
scintillation cocktail and the quantity of tritium is measured by counting the
light flashes that occur in the scintillation cocktail as the tritium decays. The
surface contamination is then calculated in units of dpm (100 cm2)–1. For liquid
measurement, a sample of the liquid to be measured is placed in the liquid
scintillation cocktail and measured. The tritium concentration of the liquid is
calculated in Bq/mL or Ci/mL. For solids measurement, a known weight of a
material is dissolved to produce a liquid and then the liquid is sampled and
measured in the scintillation counter. The quantity of tritium is then calculated
in units of in Bq/g or Ci/g of the original solid.
6.2.3.6. Portable room air monitors
Several hand held portable room air monitors are on the market, and
their capabilities and ranges vary as a function of the manufacturer and the
purpose for which they were designed. It is convenient in some activities to
have the capability to connect a small hose to the monitor so that it may be
used to detect tritium leaks around equipment.
6.2.3.7. Fixed station room air monitors
Fixed station monitors are designed to be installed in fixed locations and
to be used to monitor the room air tritium concentration. Depending upon the
manufacturer, they may have several ranges, may be equipped with one or two
alarm set points and may have audible as well as visual alarms.
6.2.3.8. Glovebox atmosphere monitors
Glovebox monitors may be open mesh or closed ionization chambers and
are designed to monitor the higher levels of tritium inside glovebox
containment systems. Ionization chambers operated in a dry glovebox
environment have a tendency to become more contaminated with tritium than
those operated in air, which can lead to false high readings and require
frequent cleaning or the adjustment of alarm set points. Gold plated and virtual
wall ionization chambers have been used to reduce this tendency for monitor
contamination.
6.2.3.9. Hood and exhaust duct air monitors
Hood and exhaust duct air monitors are similar to fixed station monitors
in range and characteristics.
69
6.2.3.10. Exhaust stack air monitors
Exhaust stack monitors are similar to fixed station air monitors, except
that they generally have larger ionization chambers to increase the sensitivity
of the monitor.
6.2.4.
Specialized instrumentation
There are many other types of specialized devices and/or instrumentation
vendors, and some may be superior to those discussed here. No endorsement of
these devices should be inferred by the reader.
6.2.4.1. Remote field tritium analysis system
The Field Deployable Tritium Analysis System (FDTAS) was developed
for the remote, in situ analysis of tritium in surface waters and groundwaters.
The system uses automated liquid scintillation counting techniques, and in
laboratory and field tests has shown sufficient sensitivity to measure tritium in
water samples at environmental levels (10 Bq/L (~270 pCi/L) for a 100 min
count) on a near real time basis. The prototype FDTAS consists of several
major components: a multi-port, fixed volume sampler; an on-line water purification system using single use tritium columns; a tritium detector employing
liquid scintillation counting techniques; and serial communications devices.
The sampling and water purification system, referred to as the autosampler, is
controlled by a logic controller preprogrammed to perform a well defined
sampling, purification and flushing protocol. The tritium analyser contains
custom software in the local computer for controlling the mixing of the purified
sample with a liquid scintillation cocktail, for counting and for flushing the cell.
An external standard is used to verify system performance and for quench
correction. All operations are initiated and monitored at the remote computer
through standard telephone line communications.
6.2.4.2. Surface activity monitor
A new surface activity monitor (SAM) for measuring tritium on metal
(electrically conducting) and non-metal (electrically non-conducting) surfaces
has been recently developed at Ontario Hydro Technologies. The monitor
detects tritium on the surface and in the near surface regions by means of
primary ionization in air due to the outward electron flux from the contaminated surface. The resulting ion pairs are measured by imposing an electric
field between the contaminated surface and a collector plate. A simple
70
theoretical model relates the total tritium concentration on the surface to the
measured current.
Experiments benchmarking the application of the surface activity
monitor on metal surfaces against independent measurement techniques of
aqueous dissolution and thermal desorption show equivalence in the total
tritium activities measured. Comparison of surface activity monitor measurements with the dry polystyrene smear protocol has shown that the two methods
are complementary. Smearing measures the activity removed by the smear
action, which can be used to infer the total activity on the surface. Surface
activity monitor measurements determine the total activity on the surface,
which can be used to infer removable activity. Ontario Power Generation has
stated that this device is the only surface monitor for tritium that provides an
absolute measurement of the total activity on metallic surfaces.
Application of the surface activity monitor on a variety of nonconducting surfaces has been demonstrated. Some of the non-conducting
surfaces examined include paper, concrete, granite and wood. Experiments are
under way to extend the database of non-conducting materials measured by
SAMs and to catalogue the associated collection efficiencies. Currently, the
SAM is commercially available in two models, QP100 and QP200, which have
measurement ranges of 0–200 nCi/cm2 and 0–200 µCi/cm2, respectively.
6.2.4.3. Breathalyser
A device undergoing development in Canada is the Scintrex tritium in
breath monitor. It is an automatic monitor dedicated to health physics and
radiation biology applications. The tritium in breath monitor measures levels of
exhaled tritium within 5 min of sampling, thus saving considerable time and
effort in the monitoring process. This rapid assessment has a sensitivity level of
5 µCi/L urine equivalent, which may be sufficient for identifying cautionary
levels of in-body tritium. Preliminary development of this equipment was
performed at AECL.
7. IMMOBILIZATION AND WASTE FORM EVALUATION
Immobilization is one of the options for dealing with radioactive waste.
The objective of immobilization is to convert radioactive waste into a stable
form, which minimizes the probability of radionuclide release to the
environment during interim storage, transport and final disposal. The
71
immobilized waste leaving the nuclear facility should therefore have such
chemical, mechanical, thermal and radiolytic stability that its integrity can be
assured over the time required for the contained radionuclides to decay to an
acceptable level. Although environmental considerations are paramount, it is
also important from an economic point of view that the process cost, as well as
the volume and weight of the immobilized waste forms produced, be as low as
possible.
Many types of solid and liquid radioactive waste are immobilized by
solidification prior to disposal. Solidification includes both fixation (chemical
and physical binding of the waste within a solidifying agent) and encapsulation
(physically surrounding the waste within an agent). It should be noted that
solidification by itself does not result in a volume reduction of the waste, and, in
fact, will result in an overall volume increase. However, solidification is quite
important in terms of immobilization of the waste for transport, storage or
disposal. The process should be reliable, easy to operate and maintain, and
should yield a product with the following properties [92]:
(a)
(b)
(c)
(d)
(e)
(f)
(g)
(h)
(i)
(j)
A monolith, with no free standing water;
Free standing without an outer container;
Little or no degradation of the product with time;
An acceptable degree of radiation damage over the design storage
period;
Compatibility between the matrix material and the solidified waste;
Sufficient mechanical resistance to compression, shock, erosion, etc.;
A low leaching rate;
No detrimental effects when exposed to fire;
Readily reproducible on an industrial scale;
A low overall production cost.
A number of immobilization technologies have been studied and
developed for waste containing 14C and tritium. The following sections describe
the immobilization techniques that have been developed and used.
7.1. IMMOBILIZATION TECHNOLOGIES FOR WASTE
CONTAINING 14C
Cements, bitumen, polymers and ceramics, including glasses and oxide
ceramics, play an important role in the immobilization of radioactive waste.
However, some of these methods are costly, are difficult to operate and
produce products that may be combustible or susceptible to radiation damage.
72
Among these methods, cementation has the advantage of being a simple, low
temperature process utilizing inexpensive raw materials.
7.1.1.
Cementation
Carbon-14 is usually removed from gaseous and liquid streams as
carbonate MeCO3 (Me = Ca, Sr, Ba), typically calcium carbonate, CaCO3 or
barium carbonate, BaCO3, as described in Section 5. However, one disadvantage of CaCO3 and BaCO3 is their high solubility in low pH media [93]. At
a given calcium activity, the equilibrium total carbonate concentration in a
solution is about 1000 times higher at pH7 than at pH10. The alkalinity and
high calcium content of Portland cement help to minimize this problem and
make cement an appropriate medium for immobilizing these carbonates [94].
Cement compounds have a low solubility, a good thermal stability and are
easy to prepare; for example, slurries of carbonate from the double alkali
process can be directly incorporated into cement after settling, without the
need to obtain the carbonate in a dry condition. The carbonate products from
calcium hydroxide or barium hydroxide processes (e.g. a dry bed column or
canister) can be incorporated into a grout to form a stable matrix. Alternatively, a cement or grout-like solution can be pumped into each canister and/or
the entire canister can be encapsulated in cement. Various assessment studies
have envisaged the use of cement waste forms [95–99].
Leaching data for cement composites of CaCO3, SrCO3, BaCO3, PbCO3
and K2CO3 have been reported [99]. Differences in leaching behaviour
between these cemented waste forms are relatively small, but CaCO3 appears
to perform the best.
Relatively little work has been conducted to assess the proper loading of
a cement matrix. Since 14C is a weak beta emitter, the effects of long term
exposure to radiation, gas evolution from radiolysis and heat generation on the
stability of the concrete matrix will be minimal. Bush et al. [12] suggest a
cement loading of 30 wt% for both BaCO3 and CaCO3.
Methods for reducing the leach rate from cement blocks have been investigated. Coating the blocks with bitumen or polymers has been studied, but is
not currently used for waste containing 14C [12]. Extensive research into
cement chemistry has been carried out in order to improve cement matrix
characteristics.
7.1.2.
Other methods
Incorporating waste containing 14C into bitumen or into certain
proprietary polymers is an alternative to the use of cement. However, the long
73
term stability of an organic matrix is questionable. No studies have been made
of these possibilities for carbonate waste, but the alternative materials would
have to offer sufficient overall advantages to outweigh the lower cost of
cement. Incorporation of carbon instead of carbonate into cement or bitumen
is unlikely to present difficulties [12]. Research on more complex carbonates
that are less soluble than MeCO3 in near neutral media is ongoing. Four
additional candidate phases for 14C immobilization have been proposed by
Taylor [93]: bismuthite, (BiO)2CO3; rhodochrosite, MnCO3; hydrocerussite,
Pb3(OH)2(CO3)2; and bastnaesite, (Ce,La)(CO3)F. Bismuthite appears to have
the lowest solubility of the four candidate forms.
7.2. IMMOBILIZATION TECHNOLOGIES FOR WASTE
CONTAINING TRITIUM
Owing to the relatively short half-life of tritium (12.3 years), it is feasible
to store waste containing tritium until the tritium has decayed by a significant
amount. This approach would not require immobilization, giving the potential
for the waste to be released. However, the risks associated with this method
need to be evaluated based on experience gained in nuclear operations.
A number of compounds containing hydrogen can be utilized as immobilization or storage media for tritium, including hydrates, hydroxides, hydrides,
organic solids and absorbents. In addition, materials that adsorb compounds
containing hydrogen can also be considered [52, 100]. In most cases the
immobilization of tritium by a chemical or physical means is not sufficient to
meet the WARs for storage, transport or disposal. For long term storage or
disposal, the immobilized media must be put into some type of high integrity
containment, such as sealed stainless steel vessels. Specific requirements for
additional containment will vary with the waste form and the storage or
disposal site.
The factors important to the evaluation of tritium immobilization or
storage media are complex and must be considered together in determining the
economic and environmental consequences of a particular choice. It is unlikely
that only one material will best meet all the criteria and requirements, and the
particular waste composition may strongly influence the choice of compound.
Heat generation by tritiated waste is of little concern, owing to the low
energy associated with tritium decay (18.1 keV), which produces 2.89 ×
_
10 15 W/Bq of tritium. Similarly, the rate of 3He production is generally small.
However, container pressurization due to 3He production and composite
radiolysis should be considered for high level tritiated waste.
74
7.2.1.
Drying agents
Many types of drying agent may be suitable for tritium immobilization or
storage, including silica gel, activated alumina, calcium sulphate and molecular
sieves. These drying agents are of value for the removal of water vapour from
air or other gases, owing to their ability to absorb water rapidly. For long term
storage these materials would have to be containerized to prevent contact with
water and water vapour, which would result in tritium release. All these drying
agents are stable solids, are widely used for commercial applications and their
properties have been thoroughly studied. Among these drying agents, silica gel
and molecular sieves have been found to have better characteristics than the
other agents, but short term tritium exchange rates (with water) are of the
order of 10–3–10–4 per day.
Silica gel has a high water absorption capacity, being capable of absorbing
40 wt% water; however, it binds water strongly and exhibits low water vapour
pressures only at a low loading. Molecular sieves exhibit high water absorption
capacities and flatter isotherms than silica gel. Their capacities are typically
10_20%, based on dry weight, and they are resistant to regeneration by heating
and to many chemical environments.
The technology of the use of drying agents to remove HTO vapour from
a gas stream or a liquid effluent is well developed. Containerization and the
encapsulation of loaded drying agents in concrete and polymers is also
available.
7.2.2.
Hydraulic cements
Hydrated silicates, such as hydraulic cements and clays, also exhibit
desirable properties for immobilization or the storage of tritium as HTO. The
use of cement to encapsulate 3H2O can be considered to be a suitable option, as
cement is widely used for the immobilization of low and intermediate level
waste containing a wide range of radionuclides [101]. Portland cement may be
the cheapest material that will bind HTO strongly and have a moderately low
exchange rate.
The strength and chemical resistance of the hardened cement block is
dependent on the amount of water incorporated into the cement. Typical water
to cement ratios of 0.3 to 0.5 are used, depending on the type of cement and any
additives used. Initial loss of HTO from cement blocks exposed to moisture can
be high (up to 25% in the first week), although this reduces with time [102].
Although concrete is a monolithic solid, it is quite porous and thus a
gradual release of tritium can be expected. The leaching of tritium from cement
blocks in an aqueous environment has been studied. Typically, the fractional
75
release of tritium per day is of the order of 1 × 10–2 for the first month, but
decreases with time. This rate of tritium release can be decreased by the
application of coatings to the cement block or by polymer impregnation [52].
Studies have been carried out to evaluate materials added to, or coated on,
cement to improve its tritium retention [26]. An improvement in release rate is
shown for epoxy coatings, particularly when used in combination with other
materials. Cement coated with coal tar based paint or with paraffin releases
significantly less tritium than uncoated cement. In general, incorporation of
tritiated water into cement blocks without further encapsulation or additional
containerization does not appear to give useful retention times, owing to the
high release rate of tritium from a cement matrix.
Encapsulation in cement is also utilized as a means of immobilizing hulls
from spent fuel shearing and dissolution processes. These hulls may contain a
significant amount of the tritium formed in the reactor, as well as a number of
other radionuclides, such as 14C.
7.2.3.
Organic agents
Many organic compounds with carbon–tritium bonds exhibit a slow rate
of tritium exchange with the hydrogen of water, and thus exposure of the
environment may occur without a significant loss of tritium. Polymeric hydrocarbons are of particular interest, owing to their low volatility, chemical
stability and hydrophobic nature. Polymeric materials under study for tritium
fixation and storage include polyacetylene, Bakelite analogue polymers,
polyacrylonitrile, polyacrylate and polystyrene. Owing to their high cost,
polymeric materials appear to be best suited to applications involving high
level tritiated waste.
7.2.4.
Hydrides
Metal hydrides can be formed with a large number of metals, but only
certain metals have the necessary properties for the immobilization and long
term storage of tritium. The metallic hydrides of zirconium, titanium, hafnium,
yttrium, niobium, tantalum and uranium have been considered for tritium
fixation and disposal applications. A good candidate hydride must be stable
when exposed to water and air and have a low dissociation pressure, which
leaves zirconium, titanium, hafnium and yttrium as possible materials. The
formation of metal hydrides requires the generation of tritium gas (HT or T2)
to react with the metal. This can be achieved by electrolysis of the HTO.
Hydrides have been used for the temporary storage of concentrated
tritium. The use of uranium beds for trapping tritium gas and storage as UH3 is
76
currently practised. However, the pyrophoric nature of uranium hydride makes
it unsuitable for long term storage, and it is unsuitable for permanent disposal.
Titanium and zirconium hydrides have been identified as compounds that have
properties suitable for tritium immobilization: low dissolution pressures and
little reactivity with air and water at normal storage temperatures. Tritium gas
can be recovered by heating the hydride. Leach rates of tritium from these
metal hydrides have been measured over a period of 600 days [103] and have
been shown to be very low, with incremental leach rates in the range of 10–6 to
10–9 cm/day. The cumulative fractional release was less than 0.05% over
600 days.
Development of this immobilization technology has been carried out to
investigate the effect of different metal surface morphologies. The risks of
ignition of the hydrides have also been investigated.
7.3. ASSESSMENT OF IMMOBILIZED WASTE FORM AND
QUALITY CONTROL
The variations that occur in the physical and chemical properties of the
waste mean that the immobilization process must be adjusted in each case to
give a satisfactory waste form. The setting of suitability criteria and arrangment
of the properties in order of importance are difficult and depend on local
circumstances and the methods selected. It should be noted that the
performance requirements of the waste form for storage, transport and
disposal are likely to be set by the WARs.
Some important specific requirements for the properties of immobilized
waste forms and the selection of the immobilization technique are:
(a)
(b)
(c)
(d)
Thermal stability: Heat generated from radiation may lead to thermal
decomposition of the waste form during storage and disposal.
Oxidative stability: Oxidation is one of the causes of solid matrix
destruction and release of radionuclides from the matrix.
Resistance to water: Waste forms may be exposed to leaching by natural
waters of various compositions, depending on the storage and disposal
strategy adopted. Low solubility, hydrolytic stability and resistance to
leaching are important in ensuring the gradual release of radionuclides
under such conditions.
Resistance to radiolysis: In principle, radiolysis may lead to gas evolution,
swelling or deterioration of the mechanical properties of the waste form.
77
(e)
(f)
(g)
Mechanical properties: These should be adequate to ensure the integrity
of the waste form, especially during transport, storage and the act of
disposal.
Specific activity: This is an important determinant of the acceptability for
various transport, storage and disposal options.
Other factors: Compactness and low cost are generally desirable features.
Ease of preparation and low dose rates to operators are also desirable.
These requirements are general for all waste forms and have been
discussed in various publications [101, 102, 104–109].
Once the method of waste immobilization has been selected, an
assessment programme must be implemented to ensure that the waste form
will meet the criteria for storage, transport and disposal (i.e. will meet the
disposal facility WARs). It has to be emphasized that adequate waste form
characterizations need to be based on the properties relevant to storage,
transport and disposal, and that suitable test standards for such properties must
be available. A test programme should include the choice and/or development
of acceptable test procedures and the performance of verification tests on the
immobilized waste form [104]. A quality assurance programme should
accompany the production of waste forms. In addition, this assessment
programme should meet the requirements of the International Organization
for Standardization (ISO) in ISO 9000 or a local equivalent to facilitate orderly
record retention and adherence to the applicable requirements.
While radioisotopes such as 14C and tritium pose specific challenges to the
design and selection of the immobilization technology, the elements and scope
of the assessment programme measuring the relative effectiveness of such
technologies are not directly tied to the contaminant being immobilized. In
other words, the assessment of waste forms designed to immobilize 14C and
tritium is not fundamentally different from other assessments involving other
contaminants.
Waste containing short lived radionuclides such as tritium may be stored
until its activity has decayed to an insignificant level. The immobilization
methods for tritium discussed in Section 7.2 are suitable for the long term
storage of tritium if the immobilized media are contained by high integrity
containers such as sealed stainless steel vessels. In most cases this final waste
form will meet the criteria for storage, transport and disposal.
7.3.1.
Waste form characterization
In general, the final waste form should be properly characterized to
ensure that the waste properties meet the identified waste disposal facility
78
WARs prior to disposal. Waste characteristics can be established using a
combination of process knowledge and waste knowledge (i.e. properties of the
waste that are known through consideration of the waste generation process or
from previous sampling and analysis of the waste), and data can be obtained
from a direct sampling and analysis of the waste.
7.3.2.
Waste form testing
Different tests for the physical and chemical properties of a waste form
are recommended in different regulatory requirements and guidelines. Several
test methods for these properties are available, although only a few of them are
standardized and fully accepted by many countries. The recommended tests
include those for free water content, unconfined compressive strength, freeze–
thaw weathering, radiation resistance, biodegradation, wet–dry weathering,
hydraulic conductivity, bulk density, acid neutralization capacity and leach
resistance [102, 109–112]. Leach tests are now becoming standardized [113–
116] and hence a comparison between different waste forms is possible. The
results of leach tests are usually presented either as the cumulative fraction
released, as a cumulative leach rate or as diffusivities [104].
Tests for the evaluation of the long term effects from ionizing radiation,
corrosion by water and soil components inside the repository, and bacterial
attack, as well as for thermal conductivities and heat reactions, have not been
standardized and are therefore performed in different ways, making
comparison difficult. The tests used for the quality control of immobilized
waste relevant for processing, storage, transport and disposal are summarized
in Table 29 [104].
79
TABLE 29.
TESTS USED FOR THE QUALITY CONTROL OF
IMMOBILIZED WASTE RELEVANT FOR PROCESSING, STORAGE,
TRANSPORT AND DISPOSAL
Property
Test
Requirement
Remarks
Not swell
Not decompose
Screening test
Water resistance
Immersion in
distilled water
Leach resistance
Leaching of
contaminant
Mechanical strength
Compressive
strength, tensile
strength and
hardness
>50 kg/cm
ASTM International
standards [119–121]
Fall resistance
Fall from 9–14 m
height
Not break (9 m)
IAEA transport
requirements [122]
Form stability
Ring ball
penetration, cylinder
bending and hole
migration
Heat resistance
Control heating
Flammability
Heating with a flame
ASTM International
standards [123–126]
Frost resistance
Storage in freezer,
cycling from –40 to
+20°C
ASTM International
standards [127]
Structure
homogeneity
Homogeneity
Microscopy X ray
> 10%
diffraction,
autoradiography and
chemical analysis
ISO standard [128]
Long term stability
against radiation,
chemical and
bacterial attack
Structure changes,
radiolytic gas
evolution and
degradation
ISO and ASTM
International
standards [129, 130]
80
ASTM International
standards [113, 114,
117, 118]
ASTM International
standards [119–121]
800°C for half an
hour
IAEA transport
requirements [122]
8. STORAGE AND DISPOSAL
Options for the management of any waste containing 14C and tritium can
be divided into two major groups: storage and disposal. The difference between
storage and disposal is defined by the intention of the act: there is an option of
retrieval in the case of storage, but there is no intention of waste retrieval with
the disposal option.
Disposal of radioactive waste can be subdivided into discharge as an
effluent, either in an airborne or aqueous form, and disposal in a solid form.
Discharges of radioactive substances to the environment in the form of
airborne or liquid effluents are regulated by authorized discharge limits, as
discussed in Section 2. To meet or exceed the discharge criteria, some effluent
streams may require treatment, as discussed in Section 5, prior to discharge.
Waste containing short lived radionuclides such as tritium may be stored
until its activity has decayed to a negligible level. However, the half-life of 14C,
5730 years, is much longer than any reasonable surveillance period, and thus
storage of waste containing 14C can only be an interim measure prior to
disposal. Where disposal is not immediately available, waste is placed in
interim storage; for example, graphite moderators from graphite reactors
containing significant amounts of 14C are in interim storage, since no disposal
strategy has yet been finalized. Waste in interim storage must initially be
conditioned to maintain its integrity such that it can be readily retrieved when
disposal becomes available and to contain its contaminants.
The main criterion of a waste storage or disposal facility is that of
minimizing the potential for the release of contaminants from the waste to the
environment and thus minimizing exposure of humans. Each storage or
disposal facility has its own WARs for ensuring that the facility meets the
requirements of the regulatory bodies.
8.1. WASTE ACCEPTANCE REQUIREMENTS
WARs are requirements relevant to the acceptance of waste packages for
handling, processing, storage and disposal. In general, the WARs may vary
from site to site. However, they should clearly state:
(a)
(b)
(c)
(d)
The acceptable waste forms;
The waste tracking requirements;
The waste segregation requirements;
A list of excluded materials;
81
(e)
(f)
(g)
The waste package requirements;
The waste characterization requirements;
The limits on radiological and chemical contents.
Waste tracking is the process of identifying and recording where and
when waste originates at a facility and its movements until delivery to the
storage and/or disposal site. Waste segregation is the activity of separating, or
keeping separate, waste according to its radiological, chemical and/or physical
properties. A waste package is a combination that includes the waste or waste
form and any containers and internal barrier (e.g. a liner) prepared in
accordance with the requirements for handling, transport, storage and/or
disposal. Waste characterization is the determination of the physical, radiological, chemical and biological properties of the waste to establish the need for
further processing and/or to decide on the appropriate method for storage and/
or disposal.
Waste may need to be converted into a stable waste form prior to its
transport or disposal. The primary criteria for choosing a waste form are
stability with respect to the storage and/or disposal conditions, tolerance to
self-irradiation and compatibility with the contained waste. The release of
contaminants from a storage and/or disposal facility is a function of the
performance of the waste form containing the waste and of the performance of
the facility itself. The performance requirements of the waste form are likely to
be set by the WARs of the available facilities.
The requirements for waste forms adopted for storage, transport and
disposal vary significantly from country to country and can have a significant
impact on the immobilization techniques adopted. In addition, the requirement
to return waste from reprocessing operations to the country of origin means
that the requirements of several countries may need to be considered in
deciding upon an appropriate waste form and immobilization technology. The
waste form requirements and various waste forms have been discussed in detail
in various publications [101, 102, 104–108]. The important requirements
include:
(a)
(b)
(c)
(d)
(e)
(f)
82
Thermal stability;
Oxidative stability;
Resistance to water;
Resistance to radiolysis;
Mechanical properties;
Specific activity.
Disposal sites for low and intermediate level radioactive waste currently
range from near surface facilities to engineered geological repositories. The
radiological content of the waste is one of the main factors in the selection of
the waste disposal method (i.e. near surface disposal versus deep geological
emplacement). It is obvious that the cost of deep geological disposal is much
higher than that of near surface disposal. An overview of the current disposal
methods is presented in Ref. [131]. An approach for the derivation of quantitative acceptable criteria for near surface disposal facilities is discussed in Refs
[132, 133]. Some examples are given below for the limits of tritium and 14C
waste packages for near surface disposal in various countries.
8.1.1.
France
Waste management and storage in France is the responsibility of the
French radioactive waste organization, Agence nationale pour la gestion des
déchets radioactifs (ANDRA). ANDRA defines the waste classification of and
WARs for the activity, the nature of the radionuclides, the doses for transport
and the final waste form. This classification is not specified by law or
regulation. In general, radioactive waste is classified based on the disposal
option and the nature of the waste (the activity level, half-life of the main
radionuclides, etc.). Table 30 illustrates the correspondence of this classification
to the traditional waste classification of low and intermediate level waste
(LILW) and high level waste (HLW), considering also the half-life of the waste.
The WARs for shallow land disposal of FMA-VC waste (Table 31) were
established in France for radionuclide inventory [134].
Categories FA-VL and MA-VL are waste containing radionuclides with a
long half-life, which must be disposed of in a deep geological repository. The
first deep geological repository in France is expected to be opened in 2015.
Waste such as reprocessing sludge and hulls of fuel element structures falls into
this waste category. At present, waste in this category is interim stored at
nuclear sites, such as the CEA’s and Cogéma’s sites.
HA waste is waste containing alpha emitters and waste with heat
generation and will be disposed of in a deep geological repository. Waste in this
category is mainly vitrified waste generated at reprocessing plants.
In general, most waste containing tritium falls into the TFA category.
However, waste containing 14C is usually classified as FA-VC waste, owing to
its long half-life.
Since a large amount of graphite waste containing 14C has accumulated in
France, a study for a dedicated disposal repository will be carried out in the
future. Feasibility studies for using graphite incineration with a fluidized bed
combustor with a capacity of 30 kg/h for reducing the amount of graphite waste
83
TABLE 30. FRENCH WASTE CLASS MATRIX
Waste class
LILW, short lived (%)
LILW, long lived (%)
HLW (%)
TFA
100
0
0
FMA-VC
100
0
0
FA-VL
0
100
0
MA-VL
0
100
0
HA
0
0
100
Note: TFA is waste of very low activity, similar to ‘exempt waste’. FMA-VC, FA-VL and
MA-VL are LILW with different half-lives. HA corresponds to HLW.
TABLE 31. CRITERIA FOR SHALLOW LAND DISPOSAL IN FRANCE
Inventory (MBq/kg)
Alpha emitters
<3.7 per package
Annual fraction leached
<1 × 10–4
<5 × 10–2
Tritium
Beta–gamma emitters
3.7 < initial activity < 37
<1 × 10–1
have been carried out in co-operation with CEA, Electricité de France and
Framatome.
8.1.2.
Japan
Japan Nuclear Fuel Limited (JNFL) operates a shallow land disposal
facility for low level radioactive waste at Rokkasho-mura in Aomori
prefecture, 700 km north of Tokyo. The disposal facility commenced operation
in December 1992. The first phase involves the construction of eight repositories with a total capacity of 40 000 m3 of waste. Ultimately, the facility’s
capacity will be expanded to 600 000 m3.
Based on the performance assessment, total radionuclide inventory in the
first phase is limited by both maximum concentrations and total radioactivity;
these limits for tritium and 14C are shown in Table 32 [135]. The maximum
concentrations allowed for 14C and tritium are 3.07 × 1011 and 8.51 × 109 Bq/t,
respectively.
84
TABLE 32. MAXIMUM CONCENTRATION OF RADIOACTIVITY IN
WASTE AND TOTAL RADIOACTIVITY AT THE ROKKASHO-MURA
DISPOSAL FACILITY FOR THE FIRST 40 000 m3 OF WASTE
Tritium
14
C
8.1.3.
Maximum concentration (Bq/t)
Total radioactivity (Bq)
3.07 × 1011
8.51 × 1090
1.22 × 1014
3.37 × 1012
Spain
The Spanish National Company for Radioactive Waste Management
(ENRESA) is responsible for the operation of the El Cabril disposal facility,
which started operation in October 1992. Two activity levels are defined for
waste packages intended for El Cabril:
(a)
(b)
Level 1 waste includes waste with sufficient stability and having mass
activities below the following limits:
(i) Total alpha activity: 1.85 × 105 Bq/kg.
(ii) Individual beta–gamma activity, half-life more than five years: 1.85 ×
107 Bq/kg.
(iii) Total beta–gamma activity, half-life more than five years (except
tritium): 7.40 × 107 Bq/kg.
(iv) Tritium activity: 7.40 × 106 Bq/kg.
Level 2 waste includes waste that must satisfy strict stability requirements
and have mass activities equal to or above level 1 but below the following
limits:
(i) Total alpha activity: 3.70 × 106 Bq/kg.
(ii) 60Co activity: 3.70 × 108 Bq/kg.
(iii) 90Sr activity: 3.70 × 108 Bq/kg.
(iv) 137Cs activity: 3.70 × 108 Bq/kg.
The activity limits for a number of other radionuclides not mentioned
above are obtained by multiplication of the above 60Co or 137Cs activities by
scaling factors [136]. These radionuclide limits are listed in Table 33, together
with the maximum permissible radionuclide loading for the El Cabril disposal
facility [137].
85
TABLE 33. MAXIMUM CONCENTRATION OF RADIONUCLIDES IN
WASTE AND MAXIMUM PERMISSIBLE ACTIVITY OF RADIONUCLIDES AT THE EL CABRIL DISPOSAL FACILITY
Maximum concentration (Bq/t)
Tritium
14
C
59
Ni
60
Co
63
Ni
90
Sr
94
Nb
99
Tc
129
I
137
Cs
241
Pu
Each beta–gamma
(except tritium)
Total beta–gamma
Total alpha
8.1.4.
Level 1
Level 2
Total radioactivity
(Bq)
7.4 × 1099
—
—
—
—
—
—
—
—
—
—
1.85 × 1010
—
3.7 × 1099
3.2 × 1099
3.7 × 1011
1.3 × 1011
3.7 × 1011
1.3 × 1079
4.4 × 1059
7.4 × 1049
3.7 × 1011
2.7 × 1010
—
2.00 × 1014
2.00 × 1013
2.00 × 1014
2.00 × 1016
2.00 × 1015
2.00 × 1015
1.00 × 1012
3.20 × 1012
1.50 × 1011
3.70 × 1015
1.15 × 1014
—
7.40 × 1099
1.85 × 1089
—
—
—
2.70 × 1013
United Kingdom
The Drigg near surface disposal site, owned and operated by BNFL, has
been the principal national disposal site for low level radioactive waste in the
UK. The WARs for Drigg were established by BNFL to ensure that the facility
meets the requirements of the UK’s regulatory bodies. Based on the safety
assessment of the facility, the annual limits for tritium and 14C disposal are 0.05
and 10.0 TBq, respectively [138, 139].
8.1.5.
United States of America
In the USA low level radioactive waste is subdivided into class A, B and
C waste. This waste is regulated under 10 CFR Part 61, Licensing Requirements for Land Disposal of Radioactive Waste [140]. Regulation 10 CRF Part
61 refers to the disposal of waste in near surface facilities. Class B and C waste
forms or their containers are required to retain their gross physical properties
and identity for at least 300 years. Waste designated as class C must be located
so that the top of the waste is at least 5 m below the surface of the facility.
86
To determine whether a waste is class A, B or C the concentrations of
certain specific radionuclides must be considered. For long lived radionuclides
(e.g. 14C), if the waste contains the radionuclides listed in Table 34 and their
concentrations do not exceed 0.1 times the values given in the table, then the
waste is class A; for example, if the waste contains less than 2.96 × 1010 Bq/m3 of
14
C and negligible amounts of other radiological contaminants, then the waste
is class A waste. If the waste contains the radionuclides listed in Table 34 and
their concentrations do not exceed the values given in the table (e.g. less than
2.96 × 1011 Bq/m3 of 14C), then the waste is class C. If the waste contains the
radionuclides listed in Table 34 and their concentrations exceed the values
given in the table, then the waste is not generally acceptable for near surface
disposal.
For short lived radionuclides (e.g. tritium), if the waste contains the radionuclides listed in Table 35 and their concentrations do not exceed the values
given in column 1 (e.g. less than 1.48 × 1012 Bq/m3 of tritium), then the waste is
class A. If the concentrations exceed the values given in column 1, but do not
exceed the values given in column 2, then the waste is class B. If the concentrations exceed the values given in column 2, but do not exceed the values given in
column 3, then the waste is class C. If the concentrations exceed the values
given in column 3, then the waste is not generally acceptable for near surface
disposal.
If the waste contains more than one of the radionuclides listed in Tables
34 and 35, with concentrations within the prescribed limits, then Regulation 10
CFR Part 61 [140] provides a method for determining the waste classification.
Waste with concentrations greater than those permitted for class C is
called ‘greater than class C’ waste. This waste is not acceptable for near surface
disposal and must be disposed of in a geological repository.
8.2. STORAGE AND DISPOSAL OPTIONS FOR WASTE
CONTAINING 14C AND TRITIUM
8.2.1.
Storage options
Waste containing short lived radionuclides such as tritium may be stored
until its activity has decayed to a negligible level. However, storage of waste
containing 14C can only be an interim measure prior to disposal, owing to its
long half-life (5730 years). Waste in interim storage must initially be
conditioned to maintain its integrity such that it can be readily retrieved when
disposal is available and to contain its contaminants. The storage methods
mainly for waste containing tritium are summarized in Sections 8.2.1.1–8.2.1.4.
87
TABLE 34. LONG LIVED RADIONUCLIDES CITED IN REF. [140]
Concentration
14
2.96 × 1011 Bq/m3
2.96 × 1012 Bq/m3
8.14 × 1014 Bq/m3
7.40 × 109 Bq/m3
1.11 × 1011 Bq/m3
2.96 × 109 Bq/m3
3.70 × 106 Bq/kg
C
C in activated metal
59
Ni in activated metal
94
Nb in activated metal
99
Tc
129
I
Alpha emitting transuranic nuclides with
half-life greater than five years
241
Pu
242
Cm
14
1.30 × 108 Bq/kg
7.40 × 108 Bq/kg
TABLE 35. SHORT LIVED RADIONUCLIDES CITED IN 10 CFR 61 [140]
Concentration (Bq/m3)
Total of all nuclides with
less than a five year
half-life
Tritium
60
Co
63
Ni
63
N in activated metal
90
Sr
137
Cs
a
Column 1
Column 2
Column 3
2.59 × 1013
a
a
1.48 × 1012
2.59 × 1013
1.30 × 1011
1.30 × 1012
1.48 × 109
3.70 × 1010
a
a
a
a
12
2.59 × 10
2.59 × 1013
5.55 × 1012
1.63 × 1012
2.59 × 1013
2.59 × 1014
2.59 × 1014
1.70 × 1014
No limits are established for these radionuclides in class B or C waste. Practical considerations such as the effects of transport, handling and disposal will limit the concentrations of this waste. This waste shall be class B unless the concentrations of other
nuclides in this table determine the waste to be class C, independent of these nuclides.
8.2.1.1. Storage of HTO waste
Storage of HTO in tanks for several decades to allow for decay should be
a viable management method. Corrosion may be of concern in the long term
88
storage of tritium due, in part, to the presence of the radiolytic decomposition
products of water. It appears reasonable to assume that tank storage is feasible
for at least a 40 year period [141]. As a process, it would consist of piping the
HTO to corrosion resistant storage tanks for controlled storage. The tanks
would be sealed, monitored and, if necessary, provided with recombiners to
oxidize back to water any hydrogen produced by radiolysis. As a method, it has
application either for interim storage until separation, reuse or fixation can be
achieved or for long term storage (i.e. until significant radioactive decay has
occurred). Leakage of HTO is the primary hazard. Continuous radiation
monitoring will be required.
Tank storage of low specific activity tritiated heavy water has been
assumed to be impractical, owing to the economic penalties associated with the
loss of use of the deuterium oxide. Long term tank storage of high specific
(>40 TBq/m3) HTO is not recommended, as such concentrations could require
double and triple containment in tanks and associated piping because of the
increased hazards from leakage.
8.2.1.2. Storage of tritium gas
Storage of tritium gas has been carried out using appropriate gas
containers. These containers were principally designed for the transport of
tritium, so the tritium can be readily recovered. However, this method may
have an increased potential for leakage of the tritium compared with the use of
hydrides.
The radioactive decay of tritium to form 3He causes an increase in
pressure within the vessel, which the system needs to be able to cope with. The
pressure inside the container will eventually reach twice the initial value. The
potential embrittlement of the container caused by the helium is also a
potential problem and has been investigated. For an initial T2 pressure of 1.01 ×
104 kPa (100 atmospheres) and a temperature of 200°C, helium embrittlement
has been shown not to be a problem [142, 143].
8.2.1.3. Storage in engineered structures
In general, an engineered storage facility consists of steel shafts enclosed
within an underground, steel lined, reinforced concrete enclosure. This type of
facility has been used for interim (e.g. 20–50 years) storage for short lived, high
level radioactive waste and for transuranic waste. Information on various
storage facilities can be found in Refs [144, 145].
Large quantities of solid waste containing tritium have been stored in a
shallow land disposal facility (Centre Manche near La Hague) in France.
89
Owing to the degradation of the waste containers, a significant amount of
activity has been released from the waste packages (e.g. up to 7400 Bq/L of
activity has been found in drainwater samples). By analysing water samples
from the facility drainage, it was possible to locate the degraded waste
containers (i.e. concrete containers), which contained a total of 2200 TBq of
tritium. The waste was retrieved from the shallow land burial, re-conditioned in
new containers and put in interim storage. The HTO in the trench containing
approximately 4 TBq/m3 of tritium was pumped and collected in order to
minimize contamination of the environment.
It is important to store tritiated waste in leaktight containers, preferably
in stainless steel containers having an endurance life of about 100 years, to
allow the tritium activity to decay to an insignificant level. A carbon steel
container with a typical endurance life of 35 years is not recommended for
storing tritiated waste, since the endurance life is not long enough.
The immobilization methods discussed in Section 7.2 can also be used for
the long term storage of tritium. However, these methods have the potential
for release of tritium and the risks associated with the methods need to be
investigated.
8.2.1.4. Storage of irradiated graphite waste
Irradiated graphite moderators from Magnox reactors and AGRs are
considered to be problematic waste because they contain large quantities of
radionuclides (e.g. up to 1 MBq/g of tritium and 0.2 MBq/g of 14C). It has been
found that these contaminants can be released to the environment due to
sudden releases of Wigner energy, which will release a large amount of heat
and entail a high risk of fire. Wigner energy is the energy accumulated by the
graphite during irradiation in the reactor, which can be suddenly released by an
increase in temperature. This effect is important when the graphite irradiation
is performed at low temperatures (below 150°C). For safety and radiological
reasons, unprocessed graphite moderators from decommissioned reactors are
in interim storage inside the reactor containment. However, specific measures
must be implemented to ensure that no degradation of the waste occurs,
including:
(a)
(b)
90
The graphite waste should be stored between 10 and 35°C to prevent
sudden releases of Wigner energy; storing the waste at low temperatures
(i.e. <10°C) will present a high risk of brittle fractures occurring in the
metallic structures that support large quantities of graphite (up to 2500 t).
Oxidation of the graphite should be prevented.
(c)
(d)
(e)
The humidity should be kept low.
Gas exchange with the environment should be minimal.
A fire prevention system should be implemented.
Graphite waste in France will be in interim storage for 50 years, allowing
a significant activity decay of short lived radionuclides. A storage period of
120 years for graphite waste is employed in the UK.
8.2.2.
Disposal options
A variety of concepts for the disposal of radioactive waste have been
studied worldwide; these range from near surface disposal to deep geological
disposal. The main objective of a waste disposal facility is to minimize the
potential for release of contaminants from the waste to the environment. In
particular, much of the effort has been focused on long term waste storage and/
or disposal of high level and long lived radioactive waste. Some of these
concepts may also be suitable for the storage and/or disposal of waste
containing 14C and tritium.
8.2.2.1. Near surface disposal
This approach has been routinely used for the disposal of low level and
intermediate level radioactive waste for several decades. Near surface disposal
in unlined shafts and disposal in concrete lined trenches or caissons have been
used. Considerable amounts of waste containing 14C and tritium have been
stored and/or disposed of in near surface disposal facilities.
For the near surface disposal of waste containing tritium, the preferred
approach is to immobilize the waste in cement and dispose of the immobilized
waste in leaktight containers, which allow the tritium activity to decay to an
insignificant level (for 50–100 years). The tritiated waste should be disposed of
in dry and sealed containers to prevent tritium release.
Seabed disposal was considered by some countries as the preferable
option for the ultimate disposal of most nuclear waste other than high level
waste. However, major political changes have subsequently taken place, and
there is currently a moratorium on the dumping at sea of any radioactive waste.
Nevertheless, some experts believe that sea disposal can be the best practical
environmental option for the disposal of bulky low level waste arising from
decommissioning, as well as for tritiated waste [146]. The principles of
stabilizing the waste within containers, eliminating void volumes and establishing a monolithic form would be applied in the sea dumping option.
91
Reference [147] includes calculations of collective dose commitment arising
from marine pathways and finds them to be dominated by 14C.
8.2.2.2. Liquid injection into geological formations
The disposal of liquid radioactive waste by injection into deep geological
strata has been studied both for high level radioactive waste and for the
disposal of HTO [148]. The major factors determining the safety of the disposal
of liquid nuclear waste are the structure and properties of the geological
formations containing the collector beds into which the waste is injected, as
well as the physicochemical processes that occur in these collector beds and
which determine the localization of the waste components within specified
boundaries of a selected geological medium. Two principal injection concepts
have been proposed: deep well injection and hydrofracturing.
Deep well injection deposits the water through a borehole into a sealed
receiving stratum at a depth of 1000 m or more. The receiving stratum must
have porosity, an absence of faults and an excess capacity. Sandstone or
limestone rocks sealed with layers of clay and shale meet these requirements.
Safety requirements include site selection to ensure that the disposal formation
does not contain faults that would transmit water to aquifers and proper casing
of the injection pipe to avoid leakage into upper groundwater. This technique
has been developed and used for non-radioactive waste disposal, so experience
exists and extensive development is not required.
For deep well injection, the HTO would be temporarily stored in tanks at
the injection site. The HTO would be mixed with cement to form a slurry, which
would be pumped into the underground formation, where it would solidify.
This method could be used at the nuclear facility site if adequate geological
conditions existed there. Deep well injection for the disposal of waste
containing tritium has been considered in the Russian Federation, but has not
been put into practice [148].
Hydrofracturing involves the injection of liquid or slurry material at high
pressures into a geological formation so that the rock material is fractured or
shattered, which creates a semiporous region suitable for containing the waste.
This technique was studied at the ORNL and has been used for the disposal of
intermediate level radioactive waste [149]. Evidence of increased seismic
activity resulting from waste injection suggests that there may be long term
containment problems associated with this disposal concept.
92
8.2.2.3. Disposal in geological formations
The placement of radioactive waste into stable geological formations has
the potential for isolating the waste from humans for geological time periods
(i.e. hundreds of thousands of years). Consequently, this approach has been
investigated in several countries for the disposal of radioactive waste, including
high level and transuranic waste. As mentioned above, significant quantities of
14
C and tritium are captured within fuel elements. If it is assumed that spent
fuel will be disposed of in geological formations rather than being reprocessed,
a significant removal of the threat to human health that these radionuclides
represent will have been achieved.
The principal and most probable pathway that could result in human
radiation exposure due to this waste is the contamination of groundwater
aquifers following intrusion of water into the repository formation. The
estimated time for tritium to reach humans varies considerably with the
assumed leachability of the waste material, the migration distance to sources of
potable water and the porosity of the formation. Typical values for
groundwater velocities in porous aquifers (about 110 m/a) would indicate a
delay of only about 200 years following water intrusion into a repository.
Carbon-14 may move somewhat slower and be delayed about 1200–1500 years.
The absence of any appreciable radioactive decay of 14C during
groundwater transport means that primary reliance must be placed upon
maintaining the integrity of the repository formation itself over long time
periods. This will require selection of repository formations located in areas of
low seismic activity and surrounded by water impervious strata.
Several geological formations have been considered for radioactive waste
storage and/or disposal, including salt beds, shale formations and hard
bedrocks such as granite and basalt. Salt deposits have been considered for
radioactive waste repositories because the existence of the salt deposit itself
indicates that there is little groundwater movement in the formation. Salt also
has good heat conduction properties and its plasticity makes it self-sealing for
several types of deformation.
The storage of pressurized tritium gas cylinders in geological repositories
requires further research. Engineered storage facilities may prove to be
preferable, particularly in view of the relatively short storage times required for
tritium.
8.2.3.
Other options
A number of options based on the reduced volume of water from a
modified Purex process have been studied, including in situ solidification in
93
underground salt caverns and conversion of HTO to hydrogen followed by
either discharge into the atmosphere or transformation to metal hydrides.
The in situ solidification method consists of mixing cement and perhaps
other low and intermediate level waste with HTO and pumping it down into a
salt dome cavern.
Conversion of HTO to hydrogen and then either direct discharge or
fixation on metal hydrides is one option (Section 7.2.4). Discharge of tritiated
hydrogen into the atmosphere is uncertain until the fate of the released
hydrogen is more fully understood and a reliable dose assessment can be made.
Fixation on metal hydrides is expensive unless the volume is small.
8.3. STRATEGY FOR THE MANAGEMENT OF WASTE
CONTAINING 14C AND TRITIUM
Waste containing low levels of 14C and tritium that meets the WARs can
be disposed of in existing waste disposal facilities (e.g. near surface disposal).
However, for waste containing high levels of 14C and tritium, such as spent ion
exchange resins from HWRs, that exceeds the WARs for near surface disposal,
no disposal facility is available at present. Different waste management
practices are therefore required.
Waste containing high concentrations of tritium (with a half-life of
12.3 years) can be stored in high integrity containers (e.g. stainless steel
containers with a life of approximately 100 years) until its activity has decayed
to a level that meets the WARs for near surface disposal. The waste is then
disposed of in the available near surface disposal facilities.
The above storage–decay option is not, however, suitable for waste
containing high levels of 14C, owing to its long half-life (5730 years). This waste
can only be in interim storage until a suitable disposal facility becomes
available.
9. CONCLUSIONS
Carbon-14 and tritium are produced by nuclear reactions that occur
naturally in the environment, in nuclear weapon testing and in nuclear reactors.
When released to the environment these radioisotopes are distributed globally,
owing to their long half-lives and residence times in the atmosphere and hydro-
94
sphere. Since both 14C and tritium are mobile in the environment, it is
important to control their release from nuclear facilities and waste
management sites by means of appropriate operational procedures and waste
management strategies and practices.
In assessing the impact of 14C and tritium releases, the consequences of
exposures of an individual in the immediate vicinity outside the exclusion
boundary area, and of the public at large, should be considered. Appropriate
effluent and environmental monitoring programmes (e.g. for 14C and tritium)
are therefore required for each nuclear facility to ensure the protection of the
public and the environment from radioactive discharges.
Discharge limits for radioisotopes are established in most countries in
accordance with the recommendations of the ICRP and by taking into account
other country or site specific factors and requirements. These limits differ from
one site to another, depending on assumptions on the nature of the effluent and
on the environment into which the discharges are made. In an authorization,
such as a licence for a nuclear facility, in most cases tritium discharge limits are
established independently of other emission limits, such as for halogens, particulates and beta–gamma and alpha emitters, for both gaseous and liquid
effluents. Discharge limits for 14C in liquid effluents are not always established
independently, and in some countries they are grouped with other beta and
gamma emitters as a gross beta–gamma limit.
Most human-made 14C and tritium is produced as a by-product (or as a
special product) in nuclear reactor operations. Both the amounts of 14C and
tritium produced and their chemical form depend on the reactor design, as does
the subsequent behaviour of these nuclides and the pathways by which they can
be released to the environment. It has been found that the majority of these
radioisotopes are contained in gaseous waste (e.g. reactor off-gases and
dissolver off-gases at reprocessing plants) and solid waste (e.g. graphite
moderators, fuel cladding, spent ion exchange resins from HWRs and
structural materials in the reactor core). Only a small quantity of 14C is released
via liquid effluents from nuclear facilities. However, the amounts of tritium in
the liquid effluents from some nuclear facilities may be considerable. In
general, tritium discharge limits are higher for HWRs than for LWRs.
Large quantities of 14C and tritium are retained in the structural materials
of a nuclear facility, which will be handled during the facility’s decommissioning phase. The graphite waste from Magnox reactors and AGRs requires
specific considerations, since high inventories of 14C and tritium are expected in
this type of waste. Currently, this type of waste is in interim storage in France
and the UK.
95
It is important to take into consideration the potential production of 14C
and tritium at the design stage of a new type of reactor. To minimize the
production of 14C, the following activities have been identified:
(a)
(b)
(c)
Minimizing nitrogen impurities in the fuel, the moderator structures (e.g.
the graphite moderator) and the structural materials in the reactor core;
Avoiding the use of chemicals (e.g. nitrogen blanketing) that will promote
the production of 14C;
Minimizing or eliminating air ingress, especially in the moderator and
primary coolant systems, through better system design.
The following actions have been identified to minimize the production of
tritium:
(i) Improving fuel fabrication to avoid occurrences of fuel cladding failures.
(ii) Improving fuel cladding materials, such as the use of zircaloy instead of
stainless steel.
(iii) Reducing lithium impurities or contamination in the reactor coolant, the
moderator and the core structures. If LiOH is used for pH adjustment in
PWR primary coolant, enriched 7Li (99%) for LiOH should be used to
minimize tritium production.
Improved operation practices can also reduce 14C and tritium emissions;
for example, reducing the frequency of venting and purging of the cover gases
(e.g. the moderator or primary coolant) has been shown to decrease 14C and
tritium emissions. Separation or removal of 14C and tritium from waste and
effluent will be beneficial to their management. Technologies for the
separation or removal of these radionuclides from gaseous and liquid effluents
are available; some are well proven, but others are in the development stage.
Most of the available techniques for 14C removal from gaseous streams
remove 14C as CO2. The main disadvantage of these methods is that a large
amount of secondary waste is generated if the waste stream contains considerable amounts of 12CO2. Although methods for the separation of 14CO2 and
12
CO2 are available, they are relatively costly. No specific study has been
carried out for the removal of 14C from liquid streams, although most of the
conventional liquid treatment methods can remove 14C as bicarbonate and
carbonate. Since more stringent environmental regulation can be expected in
the future, specific studies for the removal of 14C from liquid waste will be
required.
The typical processes for the removal of tritium from gas streams involves
converting HT to HTO, followed by removal of the HTO. Most of the proven
96
technologies for tritium removal from liquids are for applications of the detritiation of heavy water. Tritium removal from light water, however, requires
larger stripping and recovery factors than water–heavy water upgrade. Several
other processes may be suitable for the removal of tritium from liquid water,
including CECE, Girdler sulphide and bithermal hydrogen–water exchange.
Other techniques have been developed at the laboratory scale and could be
used in the future.
Among the various binding materials, cement is the most promising and
appropriate for immobilizing waste containing 14C. Various immobilization
methods, such as the use of hydrates, hydrides and organic agents, can be used
for the long term storage of tritium. However, these immobilization media
should be contained in high integrity leaktight containers for storage, in order
to allow for the natural decay of the tritium activity for a significant period of
time (50–100 years).
Waste containing low levels of 14C and tritium that meets the WARs can
be disposed of in existing waste disposal facilities (e.g. near surface disposal
facilities). Special attention must be paid to the disposal of tritiated waste,
owing to the mobility of tritium. It is recommended to dispose of this waste in
leaktight and stable containers that can prevent the dispersal of tritium into
water or to an aquifer system. For release and dilution it is preferable to release
tritium gas rather than HTO.
For waste containing high levels of 14C and tritium, such as spent ion
exchange resins from HWRs, that exceeds the WARs for near surface disposal,
no disposal facility is available at present. Waste containing tritium can be
stored in high integrity containers (e.g. stainless steel containers with a life of
approximately 100 years) until the activity has decayed to a level that meets the
WARs for near surface disposal. However, this storage–decay option is not
suitable for waste containing high levels of 14C, owing to its long half-life (5730
years). This waste can only be in interim storage until a suitable disposal facility
(e.g. deep geological disposal) becomes available. Other disposal alternatives
under investigation may be employed in the future.
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CONTRIBUTORS TO DRAFTING AND REVIEW
Balonov, M.
International Atomic Energy Agency
Dubourg, M.
Société Carbone 14 Sarl, France
Efremenkov, V.
International Atomic Energy Agency
Gilpin, J.
Oak Ridge, United States of America
Mishin, E.
V.G. Khlopin Radium Institute, Russian
Federation
Rabun, R.
Westinghouse Savannah River Company,
United States of America
Wong, P.
Atomic Energy of Canada Limited, Canada
Wright, M.
British Nuclear Fuels, United Kingdom
Consultants Meetings
Vienna, Austria: 18–22 October 1999, 11–15 September 2000,
19–23 March 2001
109